Wastewater Treatment Engineering Ed. by Mohamed Samer

Wastewater Treatment Engineering Edited by Mohamed Samer

Wastewater Treatment Engineering Edited by Mohamed Samer

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Published: 14 October, 2015 ISBN-10: 953-51-2179-0 Спизжено у ExLib: ISBN-13: 978-953-51-2179-4

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Contents

Preface

Chapter 1 Biological and Chemical Wastewater Treatment Processes by Mohamed Samer Chapter 2 Bioremediation of Nitroaromatic Compounds by Deepak Singh, Keerti Mishra and Gurunath Ramanthan Chapter 3 Selection of Promising Bacterial Strains as Potential Tools for the Bioremediation of Olive Mill Wastewater by Daniela Campaniello, Antonio Bevilacqua, Milena Sinigaglia and Maria Rosaria Corbo Chapter 4 Gelation of Arabinoxylans from Maize Wastewater — Effect of Alkaline Hydrolysis Conditions on the Gel Rheology and Microstructure by Rita Paz-Samaniego, Elizabeth Carvajal-Millan, Francisco Brown- Bojorquez, Agustín Rascón-Chu, Yolanda L. López-Franco, Norberto Sotelo-Cruz and Jaime Lizardi-Mendoza Chapter 5 Perspectives on Biological Treatment of Sanitary Landfill Leachate by Andreja Žgajnar Gotvajn and Aleksander Pavko Chapter 6 Wet Air Oxidation of Aqueous Wastes by Antal Tungler, Erika Szabados and Arezoo M. Hosseini Chapter 7 Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios by Arshad Ashraf, Muhammad Saleem Pomee, Muhammad Munir Ahmad, Muhammad Yasir Waqar and Bashir Ahmad

Preface

This book provides useful information about bioremediation, phytoremediation, and mycoremediation of wastewater and some aspects of the chemical wastewater treatment processes, including ion exchange, neutralization, adsorption, and disinfection. Additionally, this book elucidates and illustrates the wastewater treatment plants in terms of plant sizing, plant layout, plant design, and plant location. Cutting-edge topics include wet air oxidation of aqueous wastes, biodegradation of nitroaromatic compounds, biological treatment of sanitary landfill leachate, bacterial strains for the bioremediation of olive mill wastewater, gelation of arabinoxylans from maize wastewater, and modeling wastewater evolution.

Chapter 1

Biological and Chemical Wastewater Treatment Processes Mohamed Samer Additional information is available at the end of the chapter http://dx.doi.org/10.5772/61250

Abstract This chapter elucidates the technologies of biological and chemical wastewater treatment processes. The presented biological wastewater treatment processes include: (1) bioreme‐ diation of wastewater that includes aerobic treatment (oxidation ponds, aeration lagoons, aerobic bioreactors, activated sludge, percolating or trickling filters, biological filters, ro‐ tating biological contactors, biological removal of nutrients) and anaerobic treatment (anaerobic bioreactors, anaerobic lagoons); (2) phytoremediation of wastewater that in‐ cludes constructed wetlands, rhizofiltration, rhizodegradation, phytodegradation, phy‐ toaccumulation, phytotransformation, and hyperaccumulators; and (3) mycoremediation of wastewater. The discussed chemical wastewater treatment processes include chemical precipitation (coagulation, flocculation), ion exchange, neutralization, adsorption, and disinfection (chlorination/dechlorination, ozone, UV light). Additionally, this chapter elu‐ cidates and illustrates the wastewater treatment plants in terms of plant sizing, plant lay‐ out, plant design, and plant location. Keywords: Wastewater treatment, biological treatment, chemical treatment, bioremedia‐ tion, phytoremediation, mycoremediation, vermifiltration, treatment plant

1. Introduction The chapter concerns with wastewater treatment engineering, with focus on the biological and chemical treatment processes. It aims at providing a brief and obvious description of the treatment methods, designs, schematics, and specifications. The chapter also answers an important question on how the different processes are interrelated and the correct order of these processes in relation to each other. The main objective of this work was to summarize the work of the eminent scientists in this field in order to provide a clear but concise chapter that can be used as a quick reference for environmental engineers and researchers, and to be

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effectively implemented in higher education teaching undergraduate and graduate students, as well as extension and outreach.

2. Chapter description and contents overview The chapter describes the biological and chemical wastewater treatment processes that include: a.

Bioremediation of wastewater using oxidation ponds, aeration lagoons, anaerobic lagoons, aerobic and anaerobic bioreactors, activated sludge, percolating or trickling filters, biological filters, rotating biological contactors, and biological removal of nutrients;

b.

Mycoremediation of wastewater using bioreactors;

c.

Phytoremediation of wastewater that includes: constructed wetlands, rhizofiltration, rhizodegradation, phytodegradation, phytoaccumulation, phytotransformation, and hyperaccumulators;

d.

Vermifiltration and vermicomposting;

e.

Microbial fuel cells for electricity production from wastewater;

f.

Chemical wastewater treatment processes that include: chemical precipitation, ion exchange, neutralization, adsorption and disinfection (chlorination/dechlorination, ozone, ultraviolet radiation);

g.

Wastewater treatment plants. The chapter elucidates and illustrates the plant sizing, plant layout, plant design, and plant location.

3. Overview 3.1. Wastewater treatment techniques Wastewater, or sewage, originates from human and home wastewaters, industrial wastes, animal wastes, rain runoff, and groundwater infiltration. Generally, wastewater is the flow of used water from a neighborhood. The wastewater consists of 99.9% water by weight, where the remaining 0.1% is suspended or dissolved material. This solid material is a mixture of excrements, detergents, food leftovers, grease, oils, salts, plastics, heavy metals, sands, and grits [1, 2]. Types of wastewaters include: municipal wastewater, industrial wastewaters, mixtures of industrial/domestic wastewaters, and agricultural wastewaters. Typical agricul‐ tural industries include: dairy processing industries, meat processing factories, juice and beverage industries, slaughterhouses, vegetable processing facilities, rendering plants, and drainage water of irrigation systems. Subsequent to primary treatment of wastewater, i.e., physical treatment of wastewater, it still contains large amounts of dissolved and colloidal material that must be removed before

Biological and Chemical Wastewater Treatment Processes http://dx.doi.org/10.5772/61250

discharge. The issue is how to transform the dissolved materials or particulate matters that are too little for sedimentation into larger particles to allow the separation processes to eliminate them. This can be accomplished by secondary treatment, i.e., biological treatment. The treatment of wastewater subsequent to the removal of suspended solids by microorganisms such as algae, fungi, or bacteria under aerobic or anaerobic conditions during which organic matter in wastewater is oxidized or incorporated into cells that can be eliminated by removal process or sedimentation is termed biological treatment. Biological treatment is termed secondary treatment. Chemical treatment, or tertiary treatment, using chemical materials will react with a portion of the undesired chemicals and heavy metals, but a portion of the polluting material will remain unaffected. Additionally, the cost of chemical additives and the environ‐ mental problem of disposing large amounts of chemical sludge make this treatment process deficient [1]. Alternatively, the biological treatment must be implemented. This treatment process implements naturally occurring microorganisms to transform the dissolved organic matter into a dense biomass that can be separated from the treated wastewater by the sedi‐ mentation process. In fact, the microorganisms utilize the dissolved organic matter as food for themselves, where the generated sludge will be far less for chemical treatment. In practice, therefore, secondary treatment tends to be a biological process with chemical treatment implemented for the removal of toxic compounds. 3.2. Aims of wastewater treatment The goals of treating the wastewaters are: a.

Transforming the materials available in the wastewater into secure end products that are able to be safely disposed off into domestic water devoid of any negative environmental effects;

b.

Protecting public health;

c.

Ensuring that wastewaters are efficiently handled on a trustworthy basis without annoyance or offense;

d.

Recycling and recovering the valuable components available in wastewaters;

e.

Affording feasible treatment processes and disposal techniques;

f.

Complying with the legislations, acts and legal standards, and approval conditions of discharge and disposal.

3.3. Biological treatment processes The secondary treatment can be defined as “treatment of wastewater by a process involving biological treatment with a secondary sedimentation”. In other words, the secondary treatment is a biological process. The settled wastewater is introduced into a specially designed bioreac‐ tor where under aerobic or anaerobic conditions the organic matter is utilized by microorgan‐ isms such as bacteria (aerobically or anaerobically), algae, and fungi (aerobically). The bioreactor affords appropriate bioenvironmental conditions for the microorganisms to

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reproduce and use the dissolved organic matter as energy for themselves. Provided that oxygen and food, in the form of settled wastewater, are supplied to the microorganisms, the biological oxidation process of dissolved organic matter will be maintained. The biological process is mostly carried out bacteria that form the basic trophic level (the level of an organism is the position it occupies in a food chain) of the food chain inside the bioreactor. The biocon‐ version of dissolved organic matter into thick bacterial biomass can fundamentally purify the wastewater. Subsequently, it is crucial to separate the microbial biomass from the treated wastewater though sedimentation. This secondary sedimentation is basically similar to primary sedimentation except that the sludge contains bacterial cells rather than fecal solids. The biological removal of organic matter from settled wastewater is conducted by microor‐ ganisms, mainly heterotrophic bacteria but also occasionally fungi. The microorganisms are able to decompose the organic matter through two different biological processes: biological oxidation and biosynthesis [1]. The biological oxidation forms some end-products, such as minerals, that remain in the solution and are discharged with the effluent (Eq. 1). The biosyn‐ thesis transforms the colloidal and dissolved organic matter into new cells that form in turn the dense biomass that can be then removed by sedimentation (Eq. 2). Figure 1 summarizes these processes. On the other hand, algal photosynthesis plays an important role in some cases (Figure 2). Oxidation: COHNS+O 2 +Bacteria ® CO 2 + NH 3 +Other end products

(1)

Biosynthesis: COHNS+O 2 +Bacteria ® C 5H7 NO 2

(2)

(Organic matter)

+Energy

(Organic matter)

(New cells)

3.3.1. Useful terms The following terms are the most used in biological treatment processes [2]: a.

DO: Dissolved Oxygen (mg L-1)

b.

BOD: Biochemical Oxygen Demand (mg L-1)

c.

BOD5: BOD (mg L-1), incubation at 15°C for 5 days

d.

COD: Chemical Oxygen Demand (mg L-1)

e.

CBOD: Carbonaceous BOD (mg L-1)

f.

NBOD: Nitrogenous (mg L-1)

g.

SOD: Sediment Oxygen Demand (mg L-1)

h.

TBOD: Total BOD (mg L-1)

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Figure 1. Biological synthesis and oxidation [3].

3.4. Chemical treatment processes In early wastewater treatment technologies, chemical treatment has preceded biological treatment. Recently, the biological treatment precedes chemical treatment in the treatment process. Chemical treatment is now considered as a tertiary treatment that can be more broadly defined as “treatment of wastewater by a process involving chemical treatment”. The mostly implemented chemical treatment processes are: chemical precipitation, neutralization, adsorption, disinfection (chlorine, ozone, ultraviolet light), and ion exchange.

4. Biological treatment of wastewater 4.1. Biological growth equation The biological growth can be described according to the Monod equation:

m = (lS) / ( KS + S)

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Wastewater Treatment Engineering

Figure 2. Photosynthesis and oxidation [2].

Where, μ is the specific growth rate coefficient; λ is the maximum growth rate coefficient that occurs at 0.5 μmax; S is the concentration of limiting nutrient, that is BOD and COD; and KS is the Monod coefficient [3]. Generally, the bacterial growth can be explained by the following simplified figure: Organics+Bacteria+Nutrients+Oxygen ® New Bacteria+CO 2 +H 2O+Residual Organics+Inorganics Several bioenvironmental factors affect the activity of bacteria and the rate of biochemical reactions. The most important factors are: temperature, pH, dissolved oxygen, nutrient concentration, and toxic materials. All these factors can be controlled within a biological treatment system and/or a bioreactor in order to ensure that the microbial growth is maintained under optimum bioenvironmental conditions. The majority of biological treatment systems operate in the mesophilic temperature range, where the optimal temperature ranges from 20°C to 40°C. Aeration tanks and percolating filters operate at the temperature of the wastewater that ranges from 12°C to 25°C; although in percolating filters, the air temperature and the ventilation rate may have a significant effect on heat loss. The higher temperatures increase the biological activity and metabolism, which result in increasing the substrate removal rate.

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However, the increased metabolism at the higher temperatures may lead to problems of oxygen limitations. 4.2. Bacterial kinetics The bacterial kinetics can be shown in Figures 3 and 4. The microbial growth curve that shows bacterial density and specific growth rate at the different growth phases is shown in Figure 3. The microbial growth curves that compare the total biomass and the variable biomass are shown in Figure 4.

Figure 3. Microbial growth curve [1].

4.3. Principles of biological treatment The principles of biological treatment of wastewater were stated by [3]. The following is a summary of the principles: 1.

The biological systems are very sensitive for extreme variations in hydraulic loads. Diurnal variations of greater than 250% are problematic because they will create biomass loss in the clarifiers.

2.

The growth rate of microorganisms is highly dependent on temperature. A 10°C reduction in wastewater temperature dramatically decreases the biological reaction rates to half.

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Wastewater Treatment Engineering

Figure 4. Microbial growth curves [1].

3.

BOD is efficiently treated in the range of 60 to 500 mg L-1. Wastewaters in excess of 500 mg L-1 BODs have been treated successfully if sufficient dilution is applied in the treatment process, or if an anaerobic process was implemented as a pretreatment process.

4.

The biological treatment is effective in removing up to 95% of the BOD. Large tanks are required in order to eliminate the entire BOD, which is not feasible.

5.

The biological treatment systems are unable to handle “shock loads” efficiently. Equali‐ zation is necessary if the variation in strength of the wastewater is more than 150% or if that wastewater at its peak concentration is in excess of 1,000 mg L-1 BOD.

6.

The carbon:nitrogen:phosphorus (C:N:P) ratio of wastewater is usually ideal. The C:N:P ratio of industrial wastewaters should range from 100:20:1 to 100:5:1 for a most advanta‐ geous biological process.

7.

If the C:N:P ratio of the wastewater is strong in an element in comparison to the other elements, then poor treatment will result. This is especially true if the wastewater is very strong in carbon. The wastewater should also be neither very weak nor very strong in an element; although very weak is acceptable, it is difficult to treat.

8.

Oils and solids cannot be handled in a biological treatment system because they negatively affect the treatment process. These wastes should be pretreated to remove solids and oils.

9.

Toxic and biological-resistant materials require special consideration and may require pretreatment before being introduced into a biological treatment system.

10. Although the capacity of the wastewater to utilize oxygen is unlimited, the capacity of any aeration system is limited in terms of oxygen transfer.

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4.4. Bioremediation of wastewater Bioremediation is a treatment process that involves the implementation of microorganisms to remove pollutants from a contaminated setting. Bioremediation can be defined as “treatment that implements natural organisms to decompose hazardous materials into less toxic or nontoxic materials”. Some examples of bioremediation-related technologies are phytoreme‐ diation, bioaugmentation, rhizofiltration, and biostimulation. The microorganisms imple‐ mented to carry out the bioremediation are called bioremediators. However, some pollutants are not easily removed or decomposed by bioremediation. For example, heavy metals such as lead and cadmium are not eagerly captured by bioremediators. Example of bioremediation: fish bone char has been shown to bioremediate small amounts of cadmium, copper, and zinc. The bioremediation of wastewater can be achieved by autotrophs or heterotrophs. A hetero‐ troph is an organism that is unable to fix carbon and utilizes organic carbon for its growth. Heterotrophs are divided based on their source of energy. If the heterotroph utilizes light as its source of energy, then it is considered a photoheterotroph. If the heterotroph utilizes organic and/or inorganic compounds as energy sources, it is then considered a chemoheterotroph. Autotrophs, such as plants and algae, that are able to utilize energy from sunlight are called photoautotrophs. Autotrophs that utilize inorganic compounds to produce organic com‐ pounds such as carbohydrates, fats, and proteins from inorganic carbon dioxide are called lithoautotrophs. These reduced carbon compounds can be utilized as energy sources by autotrophs and provide the energy in food consumed by heterotrophs. Over 95% of all organisms are heterotrophic. 4.4.1. Aerobic treatment Aeration has been used to remove trace organic volatile compounds (VOCs) in water. It has also been employed to transfer a substance, such as oxygen, from air or a gas phase into water in a process called “gas adsorption” or “oxidation”, i.e., to oxidize iron and/or manganese. Aeration also provides the escape of dissolved gases, such as CO2 and H2S. Air stripping has been also utilized effectively to remove NH3 from wastewater and to remove volatile tastes and other such substances in water [2]. Samer [4] and Samer et al. [5] mentioned that aerobic treatment with biowastes is effective in reducing harmful gaseous emissions as greenhouse gases (CH4 and N2O) and ammonia. 4.4.1.1. Oxidation ponds Oxidation ponds (Figure 5) are aerobic systems where the oxygen required by the heterotro‐ phic bacteria (a heterotroph is an organism that cannot fix carbon and uses organic carbon for growth) is provided not only by transfer from the atmosphere but also by photosynthetic algae. The algae are restricted to the euphotic zone (sunlight zone), which is often only a few centimeters deep. Ponds are constructed to a depth of between 1.2 and 1.8 m to ensure maximum penetration of sunlight, and appear dark green in color due to dense algal devel‐

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opment. Samer [6] and Samer et al. [7] illustrated the structures and constructions of the aerobic treatment tanks and the used building materials.

Figure 5. Aerobic system/oxidation pond [1].

In oxidation ponds, the algae use the inorganic compounds (N, P, CO2) released by aerobic bacteria for growth using sunlight for energy. They release oxygen into the solution that in turn is utilized by the bacteria, completing the symbiotic cycle. There are two distinct zones in facultative ponds: the upper aerobic zone where bacterial (facultative) activity occurs and a lower anaerobic zone where solids settle out of suspension to form a sludge that is degraded anaerobically. 4.4.1.2. Aeration lagoons Aeration lagoons are profound (3–4 m) compared to oxidation ponds, where oxygen is provided by aerators but not by the photosynthetic activity of algae as in the oxidation ponds. The aerators keep the microbial biomass suspended and provide sufficient dissolved oxygen that allows maximal aerobic activity. On the other hand, bubble aeration is commonly used where the bubbles are generated by compressed air pumped through plastic tubing laid through the base of the lagoon. A predominately bacterial biomass develops and, whereas there is neither sedimentation nor sludge return, this procedure counts on adequate mixed liquor formed in the tank/lagoon. Therefore, the aeration lagoons are suitable for strong but degradable wastewater such as wastewaters of food industries. The hydraulic retention time (HRT) ranges from 3 to 8 days based on treatment level, strength, and temperature of the influent. Generally, HRT of about 5 days at 20°C achieves 85% removal of BOD in household wastewater. However, if the temperature falls by 10°C, then the BOD removal will decrease to 65% [1].

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4.4.2. Anaerobic treatment The anaerobic treatments are implemented to treat wastewaters rich in biodegradable organic matter (BOD >500 mg L-1) and for further treatment of sedimentation sludges. Strong organic wastewaters containing large amounts of biodegradable materials are discharged mainly by agricultural and food processing industries. These wastewaters are difficult to be treated aerobically due to the troubles and expenses of fulfillment of the elevated oxygen demand to preserve the aerobic conditions [1]. In contrast, anaerobic degradation occurs in the absence of oxygen. Although the anaerobic treatment is time-consuming, it has a multitude of advan‐ tages in treating strong organic wastewaters. These advantages include elevated levels of purification, aptitude to handle high organic loads, generating small amounts of sludges that are usually very stable, and production of methane (inert combustible gas) as end-product. Anaerobic digestion is a complex multistep process in terms of chemistry and microbiology. Organic materials are degraded into basic constituents, finally to methane gas under the absence of an electron acceptor such as oxygen [8]. The basic metabolic pathway of anaerobic digestion is shown in Figures 6 and 7. To achieve this pathway, the presence of very different and closely dependent microbial population is required.

Figure 6. Steps of the anaerobic digestion process [8].

Suitable wastewaters include livestock manure, food processing effluents, petroleum wastes (if the toxicity is controlled), and canning and dyestuff wastes where soluble organic matters are implemented in the treatment. Most anaerobic processes (solids fermentation) occur in two predetermined temperature ranges: mesophilic or thermophilic. The temperature ranges are 30–38oC and 38–50oC, respectively [3]. In contrast to aerobic systems, absolute stabilization of organic matter is not achievable under anaerobic conditions. Therefore, subsequent aerobic treatment of the anaerobic effluents is usually essential. The final waste matter discharged by the anaerobic treatment includes solubilized organic matter that is acquiescent to aerobic

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Figure 7. Major steps in anaerobic decomposition [1].

treatment demonstrating the possibility of installing collective anaerobic and aerobic units in series [1]. 4.4.2.1. Anaerobic digesters Samer [9] elucidated and illustrated the structures and constructions of the anaerobic digesters and the used building materials. Samer [10] developed an expert system for planning and designing biogas plants. Figures 8 to 13 show different types of anaerobic digesters. While Figures 14 and 15 show some industrial applications. Table 1 shows the advantages and disadvantages of anaerobic treatment compared to aerobic treatment.

Figure 8. Most commonly used anaerobic reactor types: (A) Completely mixed anaerobic digester, (B) UASB reactor, (C) AFB or EGSB reactor, and (D) Upflow AF [8].

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Figure 9. Single-stage conventional anaerobic digester [3].

Figure 10. Dual-stage high rate digester [3].

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Figure 11. Schematic representation of digester types. Flow-through (A–B) and contact systems (C–F) [1].

Figure 12. The upper scheme shows a two-stage anaerobic sludge digester, while the lower scheme shows the conven‐ tional sludge digestion plant [1].

Biological and Chemical Wastewater Treatment Processes http://dx.doi.org/10.5772/61250

Figure 13. Primary digestion tank with screw mixing pump and external heater [1].

Figure 14. Wastewater treatment plant for corn processing industry [8].

Figure 15. Mass balance study for a wastewater treatment plant of the baker’s yeast industry [8].

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By definition, the anaerobic treatment is conducted without oxygen. It is different from an anoxic process, which is a reduced environment in contrast to an environment without oxygen. Both processes are anoxic, but anaerobic is an environment beyond anoxic where the oxidation reduction potential (ORP) values are highly negative. In the anaerobic process, nitrate is reduced to ammonia and nitrogen gas, and sulfate (SO32-) is reduced to hydrogen sulfide (H2S). Phosphate is also reduced because it is often transformed through the ADP–ATP chain [3].

Table 1. The advantages and disadvantages of anaerobic treatment compared to aerobic treatment [1].

4.4.2.2. Anaerobic lagoons An anaerobic lagoon is a deep lagoon, fundamentally without dissolved oxygen, that enforces anaerobic conditions. The anaerobic process occurs in deep ground ponds, and such basins are implemented for anaerobic pretreatment. The anaerobic lagoons are not aerated, heated, or mixed. The depth of an anaerobic lagoon should be typically deeper than 2.5 m, where deeper lagoons are more efficient. Such depths diminish the amount of oxygen diffused from the surface, allowing anaerobic conditions to prevail (U.S. EPA, 2002). Figures 16 to 18 show different types of anaerobic lagoons. 4.4.3. Bioreactors A bioreactor can be defined as “engineered or manufactured apparatus or system that controls the embraced or encompassed bioenvironment”. Precisely, the bioreactor is a vessel in which a biochemical process is conducted, where it involves microorganisms (e.g., bacteria, algae, fungi) or biochemical substances (e.g., enzymes) derived from such microorganisms. The

Biological and Chemical Wastewater Treatment Processes http://dx.doi.org/10.5772/61250

Figure 16. Anaerobic lagoon for strong wastewater treatment, such as meat processing wastewater [1].

Figure 17. Schematic of volume fractions in anaerobic lagoon design [11].

Figure 18. Anaerobic wastewater treatment lagoon [12].

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treatment can be conducted under either aerobic or anaerobic conditions. The bioreactors are commonly made of stainless steel, usually cylindrical in shape and range in size from liters to cubic meters. The bioreactors are classified as batch, plug, or continuous flow reactors (e.g., continuous stirred-tank bioreactor). Mycoremediation is a type of bioremediation where fungi are implemented to break down the contaminants. The term “mycoremediation” refers particularly to the implementation of fungal “mycelia” in bioremediation. The principal role of fungi in the ecological system is the breakdown of pollutants, which is performed by the mycelium. The mycelium, the vegetative part of a fungus, secretes enzymes and acids that biodegrade lignin and cellulose that are the main components of vegetative fibers. Lignin and cellulose are organic compounds composed of long chains of carbon and hydrogen, and therefore they are structurally similar to several organic pollutants. One key issue is specifying the right fungus to break down a determined pollutant. Similarly, mycofiltration is a process that uses fungal mycelia to filter toxic com‐ pounds from wastewater. In an experiment, wastewater contaminated with diesel oil was inoculated with mycelia of oyster mushrooms. One month later, more than 93% of many of the polycyclic aromatic hydrocarbons (PAH) had been reduced to non-toxic components in the mycelial-inoculated samples. The natural microbial community participates with the fungi to break down contaminants, eventually into CO2 and H2O. Wood-degrading fungi are particularly effective in breaking down aromatic pollutants (toxic components of petroleum), as well as chlorinated compounds (certain persistent pesticides). Figures 19 to 22 show different types and designs of bioreactors. 4.4.4. Activated sludge The activated sludge process is based on a mixture of thick bacterial population suspended in the wastewater under aerobic conditions. With unlimited nutrients and oxygen, high rates of bacterial growth and respiration can be attained, which results in the consumption of the available organic matter to either oxidized end-products (e.g., CO2, NO3-, SO42-, and PO43-) or biosynthesis of new microorganisms. The activated sludge process is based on five interde‐ pendent elements, which are: bioreactor, activated sludge, aeration and mixing system, sedimentation tank, and returned sludge [1]. The biological process using activated sludge is a commonly used method for the treatment of wastewater, where the running costs are inexpensive (Figure 23). However, a huge quantity of surplus sludge is produced in waste‐ water treatment plants (WWTPs) which is an enormous burden in both economical and environmental aspects. The excess sludge contains a lot of moisture and is not easy to treat. The byproducts of WWTPs are dewatered, dried, and finally burnt into ashes. Some are used in farm lands as compost fertilizer [15]. However, it is suggested that the dried byproducts of WWTPs are fed into the pyrolysis process rather than the burning process. The sludge volume index (SVI) is an estimation that specifies the tendency of aerated solids, i.e., activated sludge solids, to become dense or concentrated through the thickening process. SVI can be computed as follows: (a) allowing a mixed liquor sample from the aeration tank to sediment in 30 min; (b) determining the concentration of the suspended solids for a sample of the same mixed liquor; (c) SVI is then computed as ratio of the measured wet volume (mL/L)

Biological and Chemical Wastewater Treatment Processes http://dx.doi.org/10.5772/61250

Figure 19. A bioreactor for fungal degradation: trickle bed bioreactor [13].

of the settled sludge to the dry weight concentration of MLSS in g/L (Source: Office of Water Programs, Sacramento State, USA). During the treatment of wastewater in aeration tanks through the activated sludge process (Table 2) there are suspended solids, where the concentration of the suspended solids is termed as mixed liquor suspended solids (MLSS), which is measured in milligrams per liter (mg L-1). Mixed liquor is a mixture of raw wastewater and activated sludge in an aeration tank. MLSS consists mainly of microorganisms and non-biodegradable suspended solids. MLSS is the effective and active portion of the activated sludge process that ensures that there is adequate quantity of viable biomass available to degrade the supplied quantity of organic pollutants at any time. This is termed as Food to Microorganism Ratio (F/M Ratio) or food to mass ratio. If this ratio is kept at the suitable level, then the biomass will be able to consume high quantities of the food, which reduces the loss of residual food in the discharge. In other words, the more

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Figure 20. A bioreactor for fungal degradation: rotating disc bioreactor [13].

the biomass consumes food the lower the BOD will be in the treated effluent. It is important that MLSS eliminates BOD in order to purify the wastewater for further usage and hygiene. Raw sewage is introduced into the wastewater treatment process with a concentration of several hundred mg L-1 of BOD. The concentration of BOD in wastewater is reduced to less than 2 mg L-1 after being treated with MLSS and other treatment methods, which is considered to be safe water to use. Specification

Value

Unit

BOD-Sludge Loading

0.40

mg L-1

BOD-Volume Loading

0.20

mg L-1

MLSS

2000

mg L-1

COD of Influent

300

mg L-1

Amount of Influent

4.48

L d-1

Aeration Rate

3.00

L min-1

Table 2. Conventional activated sludge [15].

The biological treatment process is the most commonly implemented method for the treatment of domestic sewage. This method implements bacterial populations that possess superior

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Figure 21. Fluidized bed bioreactor [14].

Figure 22. Typical design of fluidized bed reactor system [1].

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Figure 23. Activated sludge [15].

sedimentation characteristics. The living microorganisms break down the organic matter in the wastewater and consequently purify the wastewater from biological waste [15]. According to [1], the main components of all activated sludge systems are: 1.

The bioreactor: it can be a lagoon, tank, or ditch. The main characteristic of a bioreactor is that it contains sufficiently aerated and mixed contents. The bioreactor is also known as the aeration tank.

2.

Activated sludge: it is the bacterial biomass inside the bioreactor that consists mostly of bacteria and other flora and microfauna. The sludge is a flocculent suspension of these microorganisms and is usually termed as the mixed liquor suspended solids (MLSS) that ranges between 2,000 and 5,000 mg L-1.

3.

Aeration and mixing system: the aeration and mixing of the activated sludge and the raw influent are necessary. While these processes can be accomplished separately, they are usually conducted using a single system of either surface aeration or diffused air.

4.

Sedimentation tank: clarification or settlement of the activated sludge discharged from the aeration tank is essential. This separates the bacterial biomass from the treated wastewater.

5.

Returned sludge: the settled activated sludge in the sedimentation tank is returned to the bioreactor to maintain the microbial population at a required concentration to guarantee persistence of treatment process.

Several parameters should be considered while operating activated sludge plants. The most important parameters are: (1) biomass control, (2) plant loading, (3) sludge settleability, and (4) sludge activity. The main operational variable is the aeration, where its major functions are: (1) ensuring a sufficient and continuous supply of dissolved oxygen (DO) for the bacterial population, (2) keeping the bacteria and the biomass suspended, and (3) mixing the influent wastewater with the biomass and removing from the solution the excessive CO2 resulting from oxidation of organic matter [1].

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Figure 24. Conventional activated sludge plant [3].

Figure 25. Complete mix plant [3].

Figure 26. Contact stabilization plant [3].

There are several types of activated sludge processes, e.g., conventional activated sludge plant (Figure 24), complete mix plant (Figure 25), contact stabilization plant (Figure 26), and step aeration plant (Figure 27). Figure 28 shows the food pyramid that represents the feeding relationships within the activated sludge process.

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Figure 27. Step aeration plant [3].

Figure 28. Food pyramid illustrating the feeding relationships within the activated sludge process [1].

4.4.5. Biological filters The main systems of operation of biological filters are: (a) single filtration, (b) recirculation, (c) ADF, and (d) two-stage filtration with high-rate primary biotower (Figure 29). There are several types of biological filters, for example, submerged aerated filters that are widely known as biological aerated filters (BAFs) and are the commonly implemented design (Figure 30), and the percolating (trickling) filters (Figure 31). The BAFs implement either the sunken granular

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media with upward (Figure 30a) or download (Figure 30b) flows, or floating granular media with upward flow (Figure 30c), which is the most common design of BAFs. In order to compare the biological filters and the activated sludge systems (Figures 31 and 32), the comparison is based on the oxidation that can be accomplished by three processes: 1.

Spreading the wastewater into a thin film of liquid with a large surface area, consequently the required oxygen can be supplied by gaseous diffusion, which is the case of the percolating filters.

2.

Aerating the wastewater by pumping air in the form of bubbles or stirring forcefully, which is the case of the activated sludge process.

3.

Implementing algae to produce oxygen by photosynthesis, which is the case of the stabilization ponds.

Figure 29. The main systems of operation of biological filters [1].

4.4.6. Rotating biological contactors The rotating biological contactors (RBC) system (Figure 33) can be implemented to amend and improve the available treatment processes as the secondary or tertiary treatment processes. The RBC is successfully implemented in all three steps of the biological treatment, which are BOD5 removal, nitrification, and denitrification. The process is a fixed-biofilm of either aerobic or anaerobic biological treatment system for removal of nitrogenous and carbonaceous compounds from wastewater (Figure 34). The RBC installations (Figure 35) were designed for removal of BOD5 or ammonia nitrogen (NH3-N), or both, from wastewater [1, 2].

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  (a)   

  (b)   

    (c)  Figure 30. Biological aerated filters [1].

Figure 30. Biological aerated filters [1].

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Figure 31. Relationship between the natural bacterial populations in rivers and the development of (A) trickling (per‐ colating) filter and (B) activated sludge system [1].

Figure 32. Comparison of the food chain pyramids for biological filters and activated sludge systems [1].

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Figure 33. Schematic diagram of air-drive RBC [2].

The RBC consists of media, shaft, drive, bearings, and cover (Figure 34). The RBC hardware consists of a large diameter and closely spaced circular plastic media that is mounted on a horizontal shaft supported by bearings and is slowly rotated by an electric motor. The plastic media are made of corrugated polystyrene or polyethylene material with different designs, dimensions, and densities. The model designs are based on increasing surface area and firmness, allowing a winding wastewater flow path and stimulating air turbulence [1, 2]. 4.4.7. Biological removal of nutrients 4.4.7.1. Biological phosphorous removal It is widely agreed that microorganisms utilize acetate and fatty acids to accumulate poly‐ phosphates as poly-β-hydroxybutyrate, which is an acid polymer. The precise mechanism is based on the production and regeneration of adenosine diphosphate (ADP) within the bacteria, and it involves the adenosine triphosphate (ATP). Phosphate removal requires true anaerobic conditions, which occur only when there is no other oxygen donor [3]. Figure 36 shows a phosphate removal process. This process needs long narrow tanks for maintenance of plug flow. 4.4.7.2. Biological removal of nitrogen The nitrification and denitrification processes are responsible for N2O production (Figure 37). Figure 38 shows a nitrification/denitrification system for biological removal of nitrogen.

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Figure 34. Mechanism of attached growth media in an RBC system [2].

Figure 35. RBC system [1].

4.4.8. Phytoremediation Phytoremediation is a treatment process that solves environmental problems by implementing plants that abate environmental pollution without excavating the pollutants and disposing them elsewhere. Phytoremediation is the abatement of pollutant concentrations in contami‐ nated soils or water using plants that are able to accumulate, degrade, or eliminate heavy metals, pesticides, solvents, explosives, crude oils and its derivatives, and a multitude of other

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Figure 36. Phosphate removal process [3].

Figure 37. Schematic illustration of nitrification and denitrification processes that are responsible for N2O release [16].

Figure 38. Nitrification/denitrification system for biological removal of nitrogen [3].

contaminants and pollutants from water and soils. Figures 39 through 44 show the designs of constructed wetlands where the phytoremediation takes place. The incorporation of heavy metals, such as mercury, into the food chain may be a deteriorating matter. Phytoremediation is useful in these situations, where natural plants or transgenic plants are able to phytodegrade and phytoaccumulate these toxic contaminants in their aboveground parts, which will be then harvested for extraction. The heavy metals in the harvested biomass can be further concentrated by incineration and recycled for industrial implementa‐ tion. Rhizofiltration is a sort of phytoremediation that involves filtering wastewater through

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Figure 39. Cross-sectional view of a typical subsurface flow constructed wetland [17].

Figure 40. Components of a horizontal flow reed bed: (1) drainage zone consisting of large rocks, (2) drainage tube of treated effluent, (3) root zone, (4) impermeable liner, (5) soil or gravel, (6) wastewater distribution system, and (7) reeds [1].

Figure 41. Free water surface system [18].

a mass of roots to remove toxic substances or excess nutrients. Phytoaccumulation or phy‐ toextraction implements plants or algae to remove pollutants and contaminants from waste‐

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Figure 42. Sub-surface flow system [18].

Figure 43. Components of a free water surface constructed wetland [2].

water into plant biomass that can be harvested. Organisms that accumulate over than usual amounts of pollutants from soils are termed hyperaccumulators, where a multitude of tables that show the different hyperaccumulators are available and should be referred to. In the case of organic pollutants, such as pesticides, explosives, solvents, industrial chemicals, and other xenobiotic substances, certain plants render these substances non-toxic by their metabolism and this process is called phytotransformation. In other cases, microorganisms that live in symbiosis with plant roots are able to metabolize these pollutants in wastewater. Figure 45 shows the tissues where the rhizofiltration, phytodegradation, and phytoaccumulation take place. 4.4.9. Vermifiltration Vermiculture, or worm farming, is the implementation of some species of earthworm, such as Eisenia fetida (known as red wiggler, brandling, or manure worm) and Lumbricus rubellus, to make vermicompost, also known as worm compost, vermicast, worm castings, worm humus, or worm manure, which is the end-product of the breakdown of organic matter

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Figure 44. Components of a vegetated submerged bed system [2].

Figure 45. Rhizofiltration, phytodegradation, and phytoaccumulation [19].

and considered to be a nutrient-rich biofertilizer and soil conditioner. Vermiculture can be implemented to transform livestock manure, food leftovers, and organic matters into a nutrient-rich biofertilizer. The potential use of earthworms to break down and manage sewage sludge began in the late 1970s [20] and was termed vermicomposting. The introduction of earthworms to the filtration systems, termed vermifiltration systems, was advocated by José Toha in 1992 [21]. Vermifilter is widely used to treat wastewater, and appeared to have high treatment efficiency, including synchronous stabilization of wastewater and sludge [22, 23, 24]. Vermifiltration is a feasible

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treatment method to reduce and stabilize liquid-state sewage sludge under optimal conditions [24, 25, 26]. Vermicomposting involves the joint action of earthworms and microorganisms [24, 27, 28], and significantly enhances the breakdown of sludge. Earthworms operate as mechan‐ ical blenders and by comminuting the organic matter they modify its physical and chemical composition, steadily decreasing the C:N ratio, increasing the surface area exposed to micro‐ organisms, and making it much more suitable for bacterial activity and further breakdown. Throughout the passageway is the earthworm gut, they move fragments and bacteria-rich excrements, consequently homogenizing the organic matter [29]. An intensified bacterial diversity was found in vermifilter, compared with conventional biofilter without earthworms [25]. The principle of using earthworms to treat sewage sludge is based on the perception that there is a net loss of biomass and energy when the food chain is extended [25]. Compared to other technologies of liquid-state sludge stabilization, such as anaerobic digestion and aerobic digestion [30], vermifiltration is a low-cost and an ecologically sound technique, and more suitable for sewage sludge treatment of small or developing-countries' WWTPs [23, 24, 25, 26, 31]. Figure 46 illustrates schematic diagram of a vermifilter, where the earthworms are in the filter bed.

Figure 46. Schematic diagram of a vermifilter [24].

An important application is in livestock manure treatment as shown in Figure 47, where manure is flushed out from the livestock building to a raw effluent tank then the raw effluent is screened to separate the solid waste from manure. The screened effluent is then introduced to the vermifilter to produce the vermicompost. The vermifiltered effluent is then stored in a sedimentation tank. Afterwards, the vermifiltered effluent is introduced to constructed

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wetlands where the phytoremediation process takes place. The purified water can be then used to flush the water from the livestock building.

Figure 47. Schematic diagram of a manure treatment system containing vermifiltration and phytoremediation process‐ es (Amended and redrawn from Morand et al. [32]).

4.4.10. Microbial fuel cells The microbial fuel cells (MFCs) allow bacteria to grow on the anode by oxidizing the organic matter that result in releasing electrons. The cathode is sparked with air to provide dissolved oxygen for the reaction of electrons, protons, and oxygen on the cathode, which result in completing the electrical circuit and producing electrical energy (Figure 48).

5. Chemical treatment of wastewater 5.1. Chemical precipitation The dissolved inorganic components can be removed by adding an acid or alkali, by changing the temperature, or by precipitation as a solid. The precipitate can be removed by sedimenta‐ tion, flotation, or other solid removal processes [1]. Although chemical precipitation (coagu‐ lation, flocculation) is still implemented, it is highly recommended to substitute the chemical precipitation process by phytoremediation (see previous section), where the trend is to ramp up the implementation of bioremediation and phytoremediation to reduce the use of chemi‐ cals, which is in line with the “Green Development”.

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Figure 48. Schematic diagram of the essential components of an MFC [33].

5.2. Neutralization Neutralization is controlling the pH of the wastewater whether it is acidic or alkaline to keep the pH around 7. The lack of sufficient alkalinity will require the addition of a base (Table 3) to adjust the pH to the acceptable range. Lime (CaO), calcium hydroxide (Ca(OH)2), sodium hydroxide (NaOH), and sodium carbonate (Na2CO3), also known as soda ash, are the most common chemicals used to adjust the pH [34]. The lack of sufficient acidity will require the addition of an acid to adjust the pH to the acceptable range. Sulfuric acid (H2SO4) and carbonic acid (H2CO3) are the most common chemicals used to adjust the pH. 5.3. Adsorption Adsorption is a physical process where soluble molecules (adsorbate) are removed by attachment to the surface of a solid substrate (adsorbent). Adsorbents should have an ex‐ tremely high specific surface area. Examples of adsorbents include activated alumina, clay colloids, hydroxides, resins, and activated carbon. The surface of the adsorbent should be free of adsorbate. Therefore, the adsorbent should be activated before use. A wide range of organic materials can be removed by adsorption, including detergents and toxic compounds. The most widely used adsorbent is activated carbon, which can be produced by pyrolytic carbonization of biomass [1]. Figure 49 illustrates the difference between absorption and adsorption.

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Note: a stoichiometric reaction will yield a pH of 7.0 Table 3. Neutralization: Case of acidic wastewater [34].

Activated carbon is the most implemented adsorbent and is a sort of carbon processed to be riddled with small, low-volume pores that enlarge the surface area available for adsorption. Owing to its high level of microporosity, 1 g of activated carbon has a surface area larger than 500 m2, which was determined by gas adsorption. Figure 50 shows a bed carbon adsorption unit. Note that the carbon can be regenerated by thermal oxidation or steam oxidation and reused. The adsorption capacity, one of the most important characteristics of an adsorbent, can be calculated as follows: Adsorption Capacity = Adsorbate / Adsorbent mg/g

mg

g

The factors that affect adsorption are [3]: 1.

Particle diameter: the adsorption is inversely proportional to the particle size of the adsorbent, and directly proportional to surface area.

2.

Adsorbate concentration: the adsorption is directly proportional to adsorbate concentra‐ tion.

3.

Temperature: the adsorption is directly proportional to temperature.

4.

Molecular weight: generally, the adsorption is inversely proportional to molecular weight depending upon the compound weight and configuration of pores diffusion control.

5.

pH: the adsorption is inversely proportional to pH due to surface charge.

6.

Individual properties of adsorbate and adsorbent are difficult to compare.

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7.

Iodine number: is the mass of iodine (g) that is consumed by 100 g of a substance.

Figure 49. A comparison between absorption and adsorption.

Figure 50. A bed carbon adsorption unit [35].

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5.4. Disinfection The disinfection of wastewater is the last treatment step of the tertiary treatment process. Disinfection is a chemical treatment process conducted by treating the effluent with the selected disinfectant to exterminate or at least inactivate the pathogens. The rationales behind effluent disinfection are to protect public health by exterminating or inactivating the pathogens such as microbes, viruses, and protozoan, and to meet the wastewater discharge standards. The purpose of disinfection is the protection of the microbial wastewater quality. The ideal disinfectant should have bacterial toxicity, is inexpensive, not dangerous to handle, and should have reliable means of detecting the presence of a residual. The chemical disinfection agents include chlorine, ozone, ultraviolet radiation, chlorine dioxide, and bromine [3]. 5.4.1. Chlorine Chlorine is one of the oldest disinfection agents used, which is one of the safest and most reliable. It has extremely good properties, which conform to the aspects of the ideal disinfec‐ tant. Effective chlorine disinfection depends upon its chemical form in wastewater. The influencing factors are pH, temperature, and organic content in the wastewater [3]. When chlorine gas is dissolved in wastewater, it rapidly hydrolyzes to hydrochloric acid (HCl) and hypochlorous acid (HOCl) as shown in the following chemical equation: Cl 2 +H 2 O « H + +Cl - +HOCl Free ammonia combines with the HOCl form of chlorine to form chloramines in a three-step reaction, as follows: NH 3 +HOCl ® NH 2 Cl+H 2O

NH 2Cl+HOCl ® NHCl 2 +H 2O NHCL 2 +HOCl ® NCl 3 +H 2O

Figure 51 illustrates the chlorination curve, where the formation of chloramines occurs at the breakpoint. The free chlorine residual first rises then falls until the reaction with ammonia has been completed. As additional chlorine is applied and ammonia is consumed, the chlorine residual rises again. Dechlorination is a very important process, where activated carbon, sulfur compounds, hydrogen sulfide, and ammonia can be implemented to minimize the residual chlorine in a disinfected effluent prior to discharge. Activated carbon and sulfur compounds are the most widely used [3]. The commonly used sulfur compounds are sulfur dioxide (SO2), sodium metabisulfite (NaS2O5), sodium bisulfate (NaHSO3), and sodium sulfite (Na2SO3). The dech‐ lorination reactions with the abovementioned compounds are described in the following equations:

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Figure 51. Chlorination curve [3].

SO 2 +2H 2O+Cl 2 ® H 2SO 4 +2HCl SO 2 +H 2O+HOCl ® 3H + +Cl - +SO 42-

Na 2S 2O 5 +2Cl 2 +3H 2O ® 2NaHSO 4 +4HCl NaHSO 3 +H 2O+Cl 2 ® NaHSO 4 +2HCl

5.4.2. Ozone Ozone (O3) is a very strong oxidant typically used in wastewater treatment. Ozone is able to oxidize a multitude of organic and inorganic compounds in wastewater. These reactions cause an ozone demand in the treated wastewater, which should be fulfilled throughout wastewater ozonation prior to developing an assessable residual. Ozone should be generated at the point of application for use in wastewater treatment as ozone is an unstable molecule [3]. Figure 52 illustrates the corona discharge method for making ozone. Ozone is generally formed by combining an oxygen atom with an oxygen molecule (O2) as follows: 5.4.3. Ultraviolet light Ultraviolet (UV) radiation is a microbial disinfectant that leaves no residual. It requires clear, un-turbid, and non-colored water for its implementation. The commercial UV disinfection systems use low- to medium-powered UV lamps with a wavelength of 354 nm [3]. The UV dosage can be calculated as follows:

D = I ×t

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Figure 52. Schematic drawing of corona discharge method for making ozone [3].

where, D is the UV dose (mW. s/cm2); I is the intensity (mW/cm2); and t is the exposure time (s). The advantages of UV radiation are: (1) directly effective against the DNA of many microor‐ ganisms, (2) not reactive with other forms of carbonaceous demand, and (3) provides superior bactericidal kill values while not leaving any residues. The advantage is often the disadvant‐ age, because power fluctuations, variations in hydraulic flow rates, and color or turbidity can cause the treatment to be ineffective [3]. Additionally, cell recovery and re-growth of the damaged organisms because of the inactivation of their predators and competitors has come to light. 5.5. Ion exchange Ion exchange (IX) is a reversible reaction in which a charged ion in a solution is exchanged with a similarly charged ion which is electrostatically attached to an immobile solid particle. The most common implementation of ion exchange method in wastewater treatment is for softening, where polyvalent cations (e.g., calcium and magnesium) are exchanged with sodium [36]. Practically, wastewater is introduced into a bed of resin. The resin is manufac‐ tured by converting a polymerization of organic compounds into a porous matrix. Typically, sodium is exchanged with cations in the solution [34]. The bed is shut down when it becomes saturated with the exchanged ions, where it should be regenerated by passing a concentrated solution of sodium back through the bed. Figure 53 shows the schematic illustration of organic

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cation-exchange bead. Figure 54 shows a typical ion exchange resin column. Table 4 shows the ion preference and affinity for some selected compounds.

Figure 53. Schematic illustration of organic cation-exchange bead [34].

Figure 54. Typical ion exchange resin column [37].

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Table 4. Ion preference and affinity for some selected compounds [3].

5.6. Physicochemical treatment processes The principal advanced physicochemical wastewater treatment processes are elucidated in Table 5.

Table 5. Principal advanced physicochemical wastewater treatment processes [1].

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6. Wastewater treatment plants This section shows some examples of WWTPs as shown in Figure 55 (a, b, and c) and Figure 56. On the other hand, there are some computer programs for planning and designing WWTPs (Figures 57, 58, and 59).

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Figure 55. WWTP showing: (a) layout of the plant, (b) wastewater process flow diagrams, and (c) sludge process flow diagram. Wastewater treatment: 1. Storm water overflow; 2. screening; 3. grit removal; 4. primary sedimentation; 5. aeration tanks; 6. Secondary sedimentation; 7. emergency chlorination; 8. filtration; 9. effluent outfall. Sludge treatment: 10. raw Figure11.55. WWTP showing: (a)sludge layout of the13. plant, wastewater process flow diagrams, and (c) sludge flow sludge thickeners; digestion tanks; 12. digested thickeners; power(b) house; 14. biogas storage; 15. filter press house; 16. transformer station. process A and B are administrative areas [1].  diagram. Wastewater treatment: 1. Storm water overflow; 2. screening; 3. grit removal; 4. primary sedimentation; 5.

aeration tanks; 6. Secondary sedimentation; 7. emergency chlorination; 8. filtration; 9. effluent outfall. Sludge treat‐ ment: 10. raw sludge thickeners; 11. digestion tanks; 12. digested sludge thickeners; 13. power house; 14. biogas stor‐ age; 15. filter press house; 16. transformer station. A and B are administrative areas [1].

Figure 56. Summary of the main process options commonly employed at both domestic and industrial WWTPs. Not all of these unit processes may be selected, but the order of their use remains the same [1].

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Figure 57. Screenshot of the STEADY program [3].

Figure 58. WEST software typical plant configuration [3].

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Figure 59. WEST configuration for multitank system [3].

7. Conclusions According to this study, it can be conclude that: 1.

The trend is to ramp up the implementation of bioremediation, phytoremediation, and mycoremediation to reduce the use of chemicals, which is in line with the “Green Development”.

2.

The recent developments elucidate that subsequent to the physical treatment processes (the primary treatment) the biological treatment processes come in turn as secondary treatment and precede the chemical treatment processes, which constitute the tertiary treatment.

3.

Microbial fuel cells, phytoremediation, and mycoremediation are the focus of the future development in this field.

Author details Mohamed Samer Address all correspondence to: [email protected] Cairo University, Faculty of Agriculture, Department of Agricultural Engineering, Egypt

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[26] Xing M., Zhao L., Yang J., Huang Z. & Xu Z. (2011). Distribution and transformation of organic matter during liquid-state vermiconversion of activated sludge using elemen‐ tal analysis and spectroscopic evaluation, Environ. Eng. Sci. 28 (9): 619–626. [27] Suthar S. (2008). Development of a novel epigeic–anecic-based polyculture vermireac‐ tor for efficient treatment of municipal sewage water sludge, Int. J. Environ. Waster Manage. 2: 84–101. [28] Xing M., Li X., Yang J., Huang Z. & Lu Y. (2011). Changes in the chemical characteris‐ tics of water-extracted organic matter from vermicomposting of sewage sludge and cow dung, J. Hazard. Mater. http://dx.doi.org/10.1016/j.jhazmat.2011.11.070. [29] Domínguez J. & Edwards C.A. (2004). Vermicomposting organic wastes, in: S.H. Shakir Hanna, W.Z.A. Mikhaïl (Eds.), Soil Zoology for Sustainable Development in the 21st Century, Cairo, pp. 369–395. [30] Fytili D. & Zabaniotou A. (2008). Utilization of sewage sludge in EU application of old and new methods – a review, Renew. Sust. Energ. Rev. 12: 116–140. [31] Sinha R.K., Bharambe G. & Chaudhari U. (2008). Sewage treatment by vermifiltration with synchronous treatment of sludge by earthworms: A low-cost sustainable technol‐ ogy over conventional systems with potential for decentralization, Environmentalist. 28: 409–420. [32] Morand P., Robin P., Qiu J.P., Li Y., Cluzeau D., Hamon G., Amblard C., Fievet S., Oudart D., Pain le Quere C., Pourcher A.-M., Escande A., Picot B. & Landrain B. (2009). Bio‐ mass production and water purification from fresh liquid manure by vermiculture, macrophytes ponds and constructed wetlands to recover nutrients and recycle water for flushing in pig housing. Proceedings of the International Congress “Ecological engineer‐ ing: From concepts to applications Foreword (EECA)”, 2-4 December 2009, Paris, France. [33] Logan B. (2008). Microbial Fuel Cells, Wiley. Hoboken, NJ, EEUU. 200 pp. [34] Davis M. L. (2010). Water and Wastewater Engineering: Design Principles and Practice, McGraw-Hill Companies, Inc., ISBN 978-0-07-171385-6, New York, USA. [35] Sincero, A. P. & G. A. Sincero (2003). Physical-Chemical Treatment of Water and Waste‐ water, CRC Press LLC, ISBN 1-84339-028-0, Boca Raton, Florida, USA. [36] Clifford D. A. (1999). Ion Exchange and Inorganic Adsorption, In: Water Quality and Treatment (5th Edition), R. D. Letterman (Ed.), pp. 9.1–9.91., American Water Works Association, McGraw-Hill, New York, USA. [37] U.S. EPA (1981) Development Document for Effluent Limitations: Guideline and Standards for the Metal Finishing Point Source Category, U.S. Environmental Protec‐ tion Agency, EPA/440/1-83-091, Washington, D.C., USA. [38] U.S. EPA (2002). Wastewater Technology Fact Sheet: Anaerobic Lagoons, U.S. Envi‐ ronmental Protection Agency, EPA 832-F-02-009, Washington, D.C., USA.

Chapter 2

Bioremediation of Nitroaromatic Compounds Deepak Singh, Keerti Mishra and Gurunath Ramanthan Additional information is available at the end of the chapter http://dx.doi.org/10.5772/61253

Abstract Nitroaromatics are major pollutants released in the environment during the post-in‐ dustrialization era and pose toxic effects to living organisms. Several bacterial strains have been isolated for the degradation of these nitroaromatic pollutants. Some of them have been used in field trial experiments for the removal of nitroaromatics from industrial water and groundwater. Very few bacterial pathways have been character‐ ized at genetic and molecular levels. In this review, we cover all reported degradation pathways and their gene evolution. These studies for nitroaromatics clearly indicate that most of the involved genes have evolved from preexisting enzymes by using all means of gene evolution like horizontal gene transfer, mutation, and promiscuity principle. This information has been exploited for the creation of hybrid pathways and better biocatalysts for degradation. Keywords: Nitroaromatics, biodegradation, gene evolution, oxygenase

1. Introduction Microbes are empowered to degrade environmental pollutants under different conditions to perform unusual metabolic and physiological activities [1]. The metabolic versatility of the microbes helps them utilize a range of organic/inorganic compounds for their growth. This metabolic versatility has been exploited for human benefit in various industrial products like cheese (dairy industry), insulin, antibiotics (pharmaceutical industry), etc. Microbial enzyme systems have also been used for the development of suitable biocatalysts for green chemistry applications [2]. Whole microbes have also been tested for their potential in bioremediation [3]. Organic aromatic compounds have been the main source of pollution in several water bodies [4]. Thus, a remediation and degradation study of aromatics has been a focus of intensive study. In this review, we will focus only on nitroaromatic compounds.

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1.1. Nitroaromatics Nitroaromatic compounds have at least one nitro group attached to the aromatic ring, like nitrobenzene, nitrotoluenes, nitrophenols, etc. In nature, nitroaromatic compounds are mostly found in natural products from different plants, fungi, and bacteria [5, 6]. The best known example of this is chloramphenicol, which is produced by Streptomyces venezuelae [7–9]. The role of some nitroaromatic compounds in cellular signaling has also been established. For example, 2-nitrophenol and 4-methyl-2-nitrophenol are well-known pheromones for ticks that enable them to aggregate and attach to mammals [5, 10] (Figure 1).

Figure 1. Nitroaromatics as cell signaling molecules.

In nature, several nitroaromatic compounds are produced by the incomplete combustion of fossil fuel, which releases hydrocarbons. These hydrocarbons produce nitroaromatic com‐ pounds after nitration with nitrogen dioxide present in the atmosphere. Mixtures of nitropolyaromatic hydrocarbons can be produced to form 3-nitrobiphenyl,1- and 2nitronaphthalene,3-NT and nitrobenzene by a hydroxyl radical-initiated mechanism [11–15]. 1.2. Synthetic nitroaromatic compounds (production and uses) The versatile chemistry of the nitro group ensures that nitroaromatic compounds serve as important feed stocks in different industrial processes. These compounds are commonly used in the manufacture of pharmaceuticals. For example, substituted nitrobenzenes and nitropyr‐ idines are used in the production of indoles, which are active components of several drugs and agrochemicals [16]. Paracetamol (an analgesic and antipyretic) is synthesized in a one-step reductive acetamidation from 4-nitrophenol [17]. Nitrobenzenes or halonitrobenzenes are used in the synthesis of derivatives of phenothiazines, a large class of drugs with antipsychotic properties [18, 19]. Some nitroaromatics like nitrobenzene, nitrotoluene, and nitrophenols are used in the synthesis of pesticides. For example, fluorodifen [20], bifenox, parathion [21], and carbofuran [22] are synthesized from nitrophenols. Some dinitrophenols like 2,5-dinitro-o-cresol have been used in the synthesis of herbicides, insecticides, fungicides, etc. [5, 23]. Aromatic amines are the largest feedstock group for chemical industries. It is estimated that the worldwide consumption of aniline is approximately 3 million tons [5]. This consumption grew by 7% annually till 2014 and is expected to reach 6.2 million tons in 2015 (Global Analysts report 2014 on aniline production). Aniline is used in the synthesis of drugs, pesticides, and

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explosives and used as a building block for the production of polyurethane foams, rubber, azo dyes, photographic chemicals, and varnishes [24]. Some nitroaromatics are used in the production of explosives like trinitrotoluene (TNT), which is produced by sequential nitration of toluenes. 1,3,5-Trinitrophenol (picric acid), which was prepared in 1771 as a yellow dye for fabrics [25], has also been used in explosive shells. The methyl group of TNT can be eliminated to produce trinitrobenzene (TNB), a high-energy explosive with decreased shock sensitivity [26]. 1.3. Release of nitroaromatics in the environment The estimated annual production of nitroaromatic compounds is 108 tons (http:www.ucl/agro/ abi/gebi). Chemical industries release these compounds into the environment through various sources, like the use of pesticides and the improper handling or storage of chemicals. The leakage into industrial effluents by improper disposal or through accidental spills by explosive ammunitions are commonly responsible for these compounds to find their way into the environment [27]. A recent example of an industrial accident is from China, where about 100 tons of benzene and nitrobenzene were released to Songhua River because of an accidental explosion in the factory of China National Petroleum, Jilin City, on November 13, 2005 [28]. 1.4. Toxic effects The electron withdrawing property of the nitro groups creates a charge on the molecule. It is a unique property that makes the nitro group an important functional group for different industrial synthetic processes. Simultaneously, the same property makes these molecules hazardous to the environment. This is why these compounds are given hazardous rating 3 (HR 3), where 3 shows the highest level of toxicity [29]. These are toxic to most living organisms, including humans, fishes, algae, and microorganisms [30, 31]. Their toxicity principally manifests itself due to their ability to uncouple photo or oxidative phosphorylation processes [32, 33]. Some of these compounds are also known for their ecotoxicity [34, 35], immunotoxicity [36], carcinogenicity [37], mutagenicity [38, 39], and teratogenicity [40, 41]. Some nitroaromat‐ ics are also converted into carcinogens and mutagens when metabolized by liver or intestinal microflora [42, 43]. 1.5. Treatment options 1.5.1. Physical, chemical, and physicochemical methods Different physical methods are available for the treatment of these toxic chemicals, like adsorption, incineration, photo-oxidation, hydrolysis, volatilization, etc. During adsorption, these compounds are only adsorbed on resins and separated but not destroyed completely. In incinerations, these compounds are treated at very high temperatures, which is neither cost effective nor eco-friendly because toxic NOx fumes are often released in the environment in this process. There are various reports on advanced oxidation processes (AOPs) that utilize ozone, UV radiation, hydrogen peroxide, or combinations of all these for the treatment of

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nitroaromatic compounds [44–47]. The use of hydrogen peroxide in these processes generates toxic intermediates and is therefore not cost-effective [48]. Because of limitations of these methods, biodegradation has emerged as a viable alternative. 1.5.2. Bioremediation Bioremediation involves biological agents to catalyze the degradation and transformation of recalcitrant molecules to simpler structures. Few common terms used in these processes are defined as follows: • Biodegradation is the breakdown of organic pollutants due to microbial activity. In this process, the microbe feeds on the pollutant to grow. The degradation of contaminant generates energy and microbe utilizes this energy for its growth. • Biomineralization is the process of complete biodegradation. The organic contaminants are degraded completely through a series of degradation steps and finally converted to inorganic molecules like H2O and CO2. In the process, organic molecule provides both carbon and energy to the microbe, and if organic molecule is nitroaromatic, it provides nitrogen as well. • Biotransformation is the process where in one organic molecule is modified by the action of biological agents. Sometimes, biotransformation occurs with cometabolism, where a microbe uses a substrate for its growth but transforms another substrate, which is not utilized by microbe for its growth. Microbes have been isolated from almost all the parts of biosphere. Further, their adoptability for different environmental conditions and ability to utilize even recalcitrant compounds for their foods make them suitable agents for bioremediation.

Figure 2. General principle of aerobic aromatic catabolism in bacteria. The three stages are as follows: the conversion of the growth substrate to catechol (or substituted catechol), then ring cleavage, and finally metabolism of the ring cleav‐ age product to central metabolites by either the ortho or meta pathways. (Adapted from Williams and Sayers [51])

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Biodegradation gained worldwide attention to treat toxic compounds [49]. This is because of its eco-friendliness. Supplementing the medium with readily utilizable carbon sources can enhance degradation processes. Thus, toxic intermediates are not generated and complete removal of toxic compounds is possible. During the last few decades, extensive research has been carried out for isolation of microbes with the abilities of degrading wide range of toxic nitroaromatic compounds and has been reviewed nicely [5, 49, 50]. Some common routes adopted by bacterial strains in nature during the degradation of aromatic compounds are described here. 1.6. General principles of pathways for aerobic aromatic catabolism The pathway for catabolism of aromatic compound basically has three stages (Figure 2) [51– 53]. In the first stage, the substrate undergo changes in its substituent groups by the action of mono- or dioxygenases to form catechols (or substituted catechols). The catechols then serve as substrates for the second stage of catabolism, that is, the ring opening. This process is facilitated by the action of dioxygenases, which breaks carbon–carbon bond by adding molecular oxygen and produce unsaturated aliphatic acid.

Figure 3. Central catechol pathway for the aerobic degradation of aromatics.

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There are two families of ring opening enzymes, ortho or intradioldioxygenases, which produce cis,cis muconic acid (or its derivative), and meta or extradioldioxygenases, which produce 2hydroxymuconic semialdehyde (or derivative). Both pathways (ortho and meta cleavage) are shown in Figure 3 and meta cleavage pathway for methylcatechols is shown in Figure 4. The third stage of catabolism is the conversion of the ring cleavage products into smaller com‐ pounds that can enter into central metabolic routes. In general, aromatic compounds are initially catabolized by various pathways (known as peripheral pathways), which converge on a limited number of common intermediates (catechols or its derivatives). These intermediates are further utilized by a small number of common pathways (central pathways).

Figure 4. The meta cleavage pathway of methylcatechols.

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2. Microbial degradation of nitroaromatics When a nitroaromatic compound is exposed to the environment, its biodegradation takes place either by anaerobic route or by aerobic route. Different strategies applied for the degradation of nitroaromatic compounds by bacterial strains are shown in Figure 5. (i) monooxygenase catalyzed reaction O2N Ar-OH

(ii) Dioxygenase catalyzed reaction OH Ar NO2 OH

Ar

HO NO2

OH Ar OH

Nitroaromatic compounds (v) Partial reduction of nitro groups of hydroxyl amines NHOH NO HO Ar Ar Ar HO NH3

Ar

N O-

O (iii) HydrideMeisenheimer complex

NO2 Ar

NO2 Ar+ H OH

NH2 NHOH NO Ar Ar Ar

NO2 Ar

(iv) reduction of nitro groups to amines

Figure 5. Different strategies of microbial remediation of nitroaromatics. (Adapted from Kulkarni and Chaudhari [49])

2.1. Anaerobic biodegradation of nitroaromatics In this process, nitro group is reduced to nitroso derivative, hydroxyl amines, or amines by the action of nitroreductases. The degradation of most of the (poly)nitroaromatic compounds occurs only under anaerobic conditions [49, 54, 55]. The complete mineralization of nitroaro‐ matic by a single anaerobic strain is very rare [56]. There are several reports showing that the initial step during the degradation of mono-, di-, and trinitroaromatic compounds is the reduction of nitro groups to amino groups [56–60]. 2.2. Aerobic biodegradation of nitroaromatic compounds Mono- and dinitroaromatics are mainly subjected to aerobic biodegradation and achieve to complete mineralization. Here nitroaromatics serve as source of carbon, nitrogen, and energy

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for the microbe. During the past few decades, several reports came up with isolation of microbes mineralizing different nitroaromatic compound and their degradation pathway. Few of them are extensively studied and characterized. There are different strategies in the aerobic degradation of nitroaromatics [61], which is used in nature as shown in Figure 5. 2.3. Reactions catalyzed by mono-oxygenases Mono-oxygenases are known to add single oxygen atom at a time and cause the release of nitro group. Simpson and Evans [62] reported the role of mono-oxygenase in a Pseudomonas sp., where 4-nitrophenol was converted to hydroquinone with the concomitant release of nitrite. Subsequently, Spain and Gibson (1991) reported accumulation of hydroquinone and release of nitrite from 4-nitrophenol by partially purified mono-oxygenase from a Moraxella sp. 2.4. Dioxygenase catalyzed reactions Dioxygenases are known to add two -OH groups simultaneously on the benzene ring of nitroaromatic compounds with the release of nitro group as nitrite. This type of mechanism is reported for 2,6-dinitrotoluene biotransformation by Alcaligenes eutrophus [63]. Other examples include the degradation of 2-NT [64, 65], 3-NT [66], nitrobenzene [67], and 2,4-DNT [68]. 2.5. Meisenheimer complex formation The addition of a hydride ion to the aromatic ring of nitroaromatic compound leads to the formation of a Meisenheimer complex [27]. The complex rearomatizes after the release of nitrite anion [69]. 2.6. Partial reduction of nitro groups In this mechanism, the nitro group is partially reduced to the corresponding hydroxylamine, which upon hydrolysis yields ammonia. This mechanism was reported in Comamonas acidovorans, where 4-nitrobenzoate is converted to 4-hydroxyl-aminobenzoate [69]. 2.7. Different bacterial pathways reported for the degradation of mononitrotoluenes and nitrobenzene 2.7.1. Bacterial degradation of nitrobenzene The aerobic degradation of nitrobenzene involves two major pathways: a most commonly found partial reductive pathway and a dioxygenase catalyzed pathway (Figure 6). In the oxidative degradation of NB, degradation starts with the action of nitrobenzene-1,2-dioxyge‐ nase, which converts nitrobenzene into catechol. This catechol is further cleaved by the action of catechol 2,3-dioxygenase and degraded by the meta cleavage pathway. This type of pathway is reported in Comamonas JS 765, Acidovorax sp. JS42, and Micrococcus sp. strain SMN1.

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OH OH

Aerobic Dioxygenation Route

NO2

Aerobic Partial reduction Route

Nitrobenzene

Catechol

NHOH

Hydroxyl aminobenzene

OH COOH CHO

NH2 OH

2-hydroxymuconic semialdehyde 2-aminophenol Hydrolase OH NH2

COOH COOH

COOH CHO

2-hydroxymuconate

2-amino muconic semialdehyde

O

OH COOH CH2

NH2 COOH COOH

COOH COOH

2-oxo-3-hexene-1,6-dioate

2-hydroxy-2,4pentadienoate

2-amino muconate NH3

O

H2O NH2

COOH CH2 2-oxo-4-pentenoate

COOH CH2 2-amino-2,4-pentadienoate

O COOH HO CH3 4-hydroxy-2-oxovalerate

CH3CHO + CH3COOH

Figure 6. Degradation pathways for nitrobenzene. Aerobic dioxygenation route is reported in Comamonas JS765 [67]. The aerobic partial reductive pathway is from Pseudomonas pseudoalkaligenes JS45 [61].

2.7.2. Bacterial degradation of mononitrotoluenes Different bacterial strains have been isolated from various sources, which can utilize nitroto‐ luenes as carbon source or both carbon and nitrogen source. There are several reports on different degradation pathways for mononitrotoluenes as described here.

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Nitrotoluenes may be subjected to reductive pathways (formation of aminotoluenes) [70] or partial reductive pathway, wherein a nitro group is reduced to hydroxyl amino group and finally releases ammonia [71–72]. For example, during the degradation of 4-NT by Pseudomo‐ nas sp., initially 4-NT is converted into 4-nitrobezoic acid via the formation of 4-nitrobenzyl alcohol and 4-nitrobenzaldehyde. Then the nitro group is partially reduced to hydroxylamino derivative (rather than amino derivative), which is further converted to protocatechuate without the utilization of oxygen and release of ammonia [71, 72]. This type of mechanism was first reported for the degradation of 4-nitrobenzoate by C. acidovorans NBA-10 [69]. In yet another mechanism of 4-NT metabolism by Micobacterium sp., 4-NT was first converted to 4hydoxyl aminotoluene followed by 6-amino-m-cresol. Here ammonia is released only after the ring cleavage [72]. An oxidative pathway is reported for 2-NT degradation by Acidovorax JS42 (Figure 7), wherein the initial oxidation of the aromatic ring takes place to form methylcatechols by simultaneous incorporation of both atoms of molecular oxygen and subsequent removal of the nitro group as nitrite by the action of a dioxygenase enzyme [64, 65, 73]. The role of mono-oxygenases and dioxygenases in the removal of nitro group from p-nitro‐ phenol has also been reported from a Pseudomonas sp. [74–76].

Figure 7. Degradation of 2-NT by the formation of 3-methylcatechol in Acidovorax sp. JS 42.

The toluene mono-oxygenase encoded by TOL plasmid oxidizes only the methyl group of 3NT and 4-NT but not of the 2-NT [77]. Toluene dioxygenase from Pseudomonas putida F1 and Pseudomonas sp. strain JS-150 oxidatively attacks on the methyl group of 2- and 3-NT to form corresponding nitrobenzylalcohols. The enzyme, however, attacks on the aromatic ring of 4NT to produce 2-methyl-5-nitrophenol and 3-methyl-6-nitrocatechol [81]. In both cases (either with toluene mono-oxygenase or dioxygenase as described above), the nitro group was not removed from the benzene ring and mononitrotoluene isomers did not serve as growth substrate. Degradations of monosubstituted 2-, 3-, and 4-nitrotoluenes were also reported from an adapted activated sludge system [79]. The two strains of Comamonas JS47 and JS46 capable of degrading 4-nitrobenzoate and 3nitrobenzoate respectively were immobilized on alginate beads jointly and separately, and these beads were loaded in the reactor and fed to different regimes of alternating nitrobenzoate isomer or mixed nitrobenzoate isomer. Through this experiment, it was deduced that same beads containing both strains were able to recover faster from change in input composition than different beads containing different strains [80].

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2.7.2.1. Bacterial degradation of 4-nitrotoluene Two different pathways are reported in bacterial strains (as shown in Figure 8). In aerobic methyl group oxidation (Pseudomonas sp. strain TW3), initially 4-NT is converted to 4nitrobenzoic acid, and then the nitro group is partially reduced to hydroxylamine derivative, which is further converted to protocatechuate without the utilization of oxygen but with concomitant release of ammonia [81]. In another mechanism, 4-NT degradation is followed by the formation of 4-hydroxylaminotoluene. CH2OH

CH3

Aerobic methyl group oxidation route

NO2

NO2

Aerobic partial reduction route

CH3

NHOH 4-hydroxyl-aminotoluene

4-Nitrobenzyl alcohol

CH3

CHO

OH NH2

NO2 4-nitrobenzaldehyde

2-amino-4-methylphenol

COO-

CH3 CHO COOH

NO2 4-nitrobenzoate

NH2 2-amino-5-methylmuconic semialdehyde

COO-

CH3 NO 4-nitrosobenzoate

COOH COOH NH2 2-amino-5-methyl muconic acid

COO-

NH3 HNOH 4-hydroxylaminobenzoate NH3 COO-

COOH

TCA Cycle OH OH Protocatechuate

CHO HOOC

OH

2-hydroxy-4-carboxy muconic semialdehyde

Figure 8. Degradation pathways for 4-nitrotoluene. Aerobic partial reduction rout is present in Mycobacterium sp. strain HL-4NT-1. Aerobic methyl group oxidation rout is reported in Pseudomonas strains TW3 and strain 4NT.

There are few well-characterized bacterial strains that degrades or biotransforms more than one mono-nitro compounds. 2-NT and 4-NT could transform to their corresponding amino‐

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toluenes and hydroxyl nitrotoluenes (pathways A and B, Figure 9) in P. putida OU83. Both oxidative as well as reductive attack is reported during metabolism of 3-NT [71]. Here 70% of the 3-NT was reduced to aminotoluene, whereas 30% was converted to 3-nitrophenol via the formation of 3-nitrobenzylalcohol, 3-nitrobenzaldehyde, and 3-nitrobenzoic acid. 3-Nitrophe‐ nol was further metabolized with the release of nitrite (pathway C, Figure 9). A

CH3

CH3

CH3

CH3

B

NH2

NO2

2-aminotoluene

2-nitrotoluene

NH2 4-aminotoluene

NO2 4-Nitrotoluene

CH3 HO

CH3

NO2

OH 6-hydroxy-2- nitrotoluene NO2 4-hydroxy-4-nitrotoluene C CH2OH

CH3

CHO

COOH

OH NO2

NO2 3-nitrotoluene

NO2 3-nitrobenzyl alcohol

NO2 3-nitrobenzaldehyde

NO2

NO2

3-nitrobenzoic

3-nitrophenol

CH3

NH2 3-aminotoluene

Figure 9. Degradation of mononitrotoluenes by Pseudomonas putida strain OU83 [82]. The strain converts 2- and 4-nitro‐ toluenes to corresponding aminotoluenes and hydroxyl nitrotoluenes (pathways A and B), whereas 70% of 3-NT was converted to 3-aminotoluene and 30% was degraded via the formation of the 3-nitrophenol (pathway C).

2.7.2.2. Degradation of mononitrotoluenes by Diaphorobacter sp. strain DS2 There are very few reports available on complete mineralization of mono-nitroaromatics by single bacterial strains. The isolation and characterization of three Diaphorobacter sp. strains DS1, DS2, and DS3, which are capable of mineralizing and utilizing 3-NT as the sole source of carbon, nitrogen, and energy, was reported by our group [66]. The mineralization of 3-NT by Diaphorobacter sp. strain DS2 was found by the initial reaction catalyzed by a dioxygenase with the formation of mixtures of 3- and 4-methylcatechols. These methylcatechols were further degraded by the meta cleavage pathway. This strain was able to degrade other compounds like 2-NT, 4-NT, nitrobenzene (NB), 2CNB, and 3CNB through an oxidative degradation route. Cloning and sequencing of first enzyme of the pathways showed the presence of a multicom‐ ponent dioxygenase, i.e., 3NTDO. This 3NTDO gene was found to be present on the genomic DNA of the strain on a 5-kb DNA stretch. Its subunits were identified as a reductase, a ferredoxin, an oxygenase large and small subunit, and a regulatory gene product [83]. Subunits

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of 3NTDO were individually expressed in E. coli and purified by various purification techni‐ ques. To get active enzyme, all the subunits were mixed together in a certain proportion with added NADH. Its active recombinant-reconstituted enzyme showed the conversion of nitrotoluenes and nitrobenzenes to methylcatechols and catechols as measure products with the release of nitrite [84]. Some other minor products are also formed as shown in Table 1. Products other than (methyl)catechols are dead-end products and observed in the culture broth during the degradation of its corresponding substrates with the Diaphorobacter sp. strain DS2. This strain did not grow on any dinitrotoluenes (2,4 or 2,6), but recombinant-reconstituted 3NTDO released nitrite from 2,6-dinitrotoluene.

Table 1. Products from the various substrates by recombinant 3NTDO.

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2.8. Degradation of dinitrotoluene 2.8.1. 2,4- and 2,6-dinitrotoluene Burkholderia sp. strain DNT and strain R34 mineralized 2,4-DNT with a nitrite removal pathway involving dioxygenase and mono-oxygenase enzyme. Here both nitro groups are not removed simultaneously but in a stepwise fashion. First, dinitrotoluene dioxygenase (DNTDO) attacks on the benzene ring converting DNT into 4-methyl-5-nitrocatachol (4M5NC) with the simultaneous removal of a nitrite group. Further, MNC mono-oxygenase removes another nitrite and converts the substrate into 2-hydroxy-5-methylquinone. These dioxyge‐ nase and mono-oxygenase enzymes have been cloned and characterized [68, 85]. This 2hydroxy-5-methylquinone eventually leads to the formation of 2,4,5-trihydroxytoluene, a reaction that is catalyzed by HMQ reductase. Ring fission of 2,4,5-trihydroxytoluene likely occurs at position 5,6 of the aromatic ring to yield 2,4-dihydroxy-5-methyl-6-oxo-hexa-2,4dienoic acid as ring cleavage product (Figure 10). Nishino et al. [61, 86] isolated Burkholderia cepacia strain JS850 and Hydrogenophaga paleronii strain JS863 that were able to mineralize 2,4-DNT in the same way but degraded 2,6-DNT in a different way. When 2,4- and 2,6-DNT were used as the sole source of carbon and nitrogen together, the dioxygenation of 2,6-DNT to 3-methyl-4-nitrocatachol (3M4NC) was the initial reaction, accompanied by the release of nitrite. 3M4NC was then subjected to meta-ring cleavage (Figure 11) without releasing the second nitro group prior to the ring cleavage. Although 2,4-DNT-degrading strains also could convert 2,6-DNT to 3M4NC, further catabo‐ lism was halted at the point. The pathway for 2,4-DNT degradation was different from that for 2,6-DNT degradation. In the latter case, 2-hydroxy-5-nitro-6-oxohepta-2,4-dienoic acid was the first ring fission product. How 3M4NC is converted to 2-hydroxy-5-nitro-6-oxohepta-2,4dienoic acid is unknown. In this pathway, the second nitrite group is released at the latter stages of the pathway. Position 3 methyl group appears as the determinant recognized by the initial dioxygenase to produce highly specific 3M4NC in the 2,6-DNT pathway proposed. The gene encoding this dioxygenase showed a nucleotide sequence similar to the α subunit among nitroarene dioxygenase. The genes for the initial dioxygenases involved in 2,4-DNT and 2,6-DNT degradation are all closely related, but the enzymes are produced at low constitutive levels [27, 86]. After initial dioxygenation, the two pathways appear to diverge (Figures 10 and 11). How does DNT degradation get affected by the presence of both isomers is important since 2,4-DNT and 2,6-DNT are produced in a 4:1 ratio [87]. These are often present together in munitions plant wastewater. Lendenmann and Spain [88] initially failed to observe the degradation of 2,4-DNT and 2,6-DNT simultaneously. Subsequently, an aerobic biofilm, initially fed with low concentrations of DNT mixture, was tested. These concentrations were then gradually increased and exhibited mineralization rates of 98% and 94% for 2,4- and 2,6DNT, respectively. The nitrogen was released as nitrite, reflecting oxidative bacterial activity. Isomer concentration needed to be kept below inhibitory levels as high concentrations of each isomer inhibited the degradation of the other. The simultaneous degradation of 2,4- and 2,6DNT may be unpredictable until an adapted population is established [87].

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Figure 10. 2,4-DNT metabolism pathway in Burkholderia sp. strain DNT and strain R34.

Although nitrotoluene degraders are widely distributed at contaminated sites, the contami‐ nants still persist for very long periods, leaving unanswered questions as to why biodegrada‐ tion is ineffective to remove them. Efficient anaerobic pathways for the degradation either of mono- or dinitrotoluenes are not known, and 2,3-DNT currently does not appear to be degradable [87].

Figure 11. 2,6-DNT metabolism pathway in Burkholderiacepacia strain JS850 and Hydrogenophaga paleroni strain JS863.

Bae et al. [89] found that in an anaerobic fluidized-bed granular carbon bioreactor, 2,4 DNT can be converted completely to 2,4-diamino toluene, which subsequently mineralized in batch activated sludge reactor. Paca et al. [90] took a mixture of microbes found in the mononitro‐ toluene, 2,4-dintrotoluene, and 2,6-dinitrotoluene contaminated soil. These microbes were extracted and immobilized on the packing material of the packed bed reactor (PBR). Varying concentrations of 2,4-DNT and 2,6-DNT were used. In this case, two types of packing material were used out of which the reactor packed with Poraver removed 97% DNT in 11 days and the one packed with fine clay achieved the efficiency of 78%. After 20 days, the metabolites detected were 2-amino-4-nitrotoluene and 2,4-diamino toluene. Wang et al. [91] reported that in wastewater enriched with contaminated DNT taken from Qingyang chemical industry with a DNT concentration of 3.55–95.65 mg/L, ethanol was mixed in the wastewater to act as an electron donor. The reactor in this case was made of polymethyl metacrylate containing polyurethane foams for microorganism immobilization. The microor‐ ganism used was B925. Initially, the reactors were domesticated and immobilized with

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microorganism for first 10 days, and then the whole system was further operated for 140 more days, gradually increasing the concentration of 2,4 DNT. DNT was then transformed to 2amino-4-nitrotoluene and 4-amino-2 nitrotoluene and 2,4-diaminotoluene. 2.9. Degradation of Trinitrotoluene (TNT) Trinitrotoluene (TNT) is very difficult to degrade [87]. The three nitro groups with a nucleo‐ philic aromatic ring structure make TNT vulnerable to reductive attack but resistant to oxygenase attack from aerobic organisms [92]. In most current reports, the reductive mecha‐ nism predominates in TNT degradation. New evidence indicates that TNT could be reduced by carbon monoxide dehydrogenase from Clostridium thermoaceticum [93] and by the manga‐ nese-dependent peroxidase (MnP) from the white-rot fungus Phlebia radiata [94]. Based on the discovery of pentaerythritoltetranitrate (PETN) reductase from Enterobacter cloacae PB2, French et al. [95] found that this strain could grow slowly on 2,4,6-TNT under aerobic conditions as the sole nitrogen source without the production of dinitrotoluene as an intermediate and catalyzed conversion of the TNT via a hydride–Meisenheimer complex with the nitro group released as nitrite. Bacteria basically tends to biotransform TNT to aminonitroaromatic compounds through aerobic degradation, which in many cases turn out to be the dead-end products, and this reduced dead-end products sometimes react with themselves and form azotetranitrotoluene [96]. The removal of nitro group from the ring is essential to allow the dioxygenase to act upon it. There are very rare cases where the complete usage of TNT as the sole carbon nitrogen and energy source has also been reported. In general, most bacteria are only capable of transform‐ ing TNT to other simpler and less toxic compounds. Due to the absence of oxygen in anaerobic processes, the formation of azonitrotoluene does not take place, thereby making the degradation through bacteria more feasible and efficient. Reduction product of TNT is very prominent in case of anaerobic process, which easily forms triaminotoluene, which is far less toxic and more soluble in water than TNT (Figures 12 and 13). Collie et al. [116] observed the biodegradation of TNT in the liquid phase bioreactor by four different bacterial strain having pure TNT in a liquid medium. The initial concentration of TNT (70mg/L) was periodically extracted from the bioreactor for by-product identification with the help of HPLC. The bacteria used were one strain of Enterobacter, one strain of Pseudomonad, and two strains of Alcaligenes. Two basic intermediates, i.e., 2-amino-4,6-DNT and 4-amino-2,6 DNT, were observed with all the bacteria after 12 hours. A bench-scale reactor by Cho and coworkers [114] used P. putida HK6 (collected from RDXcontaminated soils). This was used for the degradation of several nitroaromatic compounds simultaneously in one go, i.e., TNT–RDX–atrazine–simazine (TRAS). Cells were grown in a liquid media composed of 30–100 mg TNT, 5–15 mg RDX, 20–50 mg atrazine, and 5–15 mg simazine and other basal salts like K2HPO4, NaH2PO4, MgSO4 7H2O, CaCl2 2H2O, FeCl3 6H2O in appropriate quantities, and subsequent experiments on the utilization of nitroaromatic compounds were carried out using a 2.5-L bottom driving type bench-scale reactor with a water condenser at 5°C and operated at 30°C at 150 rpm. A 10% inocula of test culture was

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Figure 12. Different reductive pathways in different bacteria: (A) Clostridium acetobutylicum, Escherichia coli, Lactobacillus sp. [97]; (B) Clostridium bifermentans CYS-1 [98]; (C) Clostridium bifermentans LJP-1 [99]; (D) Disulphovibriosp strain B [100]; (E) Disulphovibrio sp. [101]; (F) Disulphovibrio sp. [102]; (G) Methanococcus sp. strain B [103]; (H) Veillonella alkales‐ cens [57].

grown with nitroaromatic compound, and the turbidity observed showed that the bacteria P. putida HK6 was able to degrade 100 mg/l TNT, 15 mg/l RDX, 50 mg/l atrazine, and simazine in 4, 24, 2, and 4 days after incubation, respectively. In this experiment, it is noteworthy that the presence of Tween-80 in the culture led to the complete degradation of TRAS compounds, whose otherwise partial degradation was TNT (80%), RDX (35%), and simazine (78%) during the incubation period [114].

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CH3 O2N

O2N

COOH NO2

NH2

NH2

HO

NH2

O2N

F

OH NH4+

CH3 NO2

HO

OH OH

OH

CH3 O2N

NO2

NH2 2-ADNT NO2

NH2

CH3 O2N

CH3 SH

F

4-ADNT H

2,4-DANT

A ,D ,F

68

CH3

CH3 NO2

O2N

NO2

Ar-N+=N-Ar'

NHOH

H,I

O NHCOCH3

O2N

NO2

NHOH Azoxynitrotoluene (unreactive polymer)

CH3 O2N

NHOH

NHOH CH3 O2N

NO2 Yellow product NO2 TNT

B,E CH3

CH3 O2N H H

_

NO2 H H

O2N J

NO2 NO2

NO2 H H Meisenheimer complex NO2 _

CH3 NO2

NO2

Figure 13. Complete representation of the aerobic pathway of TNT conversion to its products performed by the group of bacteria shown in the figure: (A) Bacillus sp., Pseudomonas aeruginosa, Staphylococcus sp. [104]; (B) Enterobacter cloacae PB2 [105]; (C) Enterobacter sp. [89]; (D) Pseudomonas aeruginosa MA101 [106]; (E) Pseudomonas florescence [107]; (F) Pseu‐ domonas florescence B3468 [108]; (G) Pseudomonas florescence [109]; (H) Pseudomonas pseudoalcaligenes [110]; (I) Rhodococcus erythropolis [111]; (J) Serratia marcensens [112].

Similarly, the degradation of some other nitroaromatic compounds was reported on the basis of field trials. A highly active microbial consortium was chosen by Oh et al. [115] for field trials to degrade the wastewater samples having 4,6-dinitro-ortho-cresol (DNOC) [116]. Fixed bed column reactors were employed to increase the volume density of active biomass and to degrade DNOC in wastewater stream. In this case, glass bead matrix was used to immobilize the bacterial colonies. The mixed bacterial culture CDNOC1-3 metabolizes DNOC with concomitant liberation of nitrate. The efficiency of the bioreactor was 86%. This was substan‐ tially higher over a batch culture mode (60–65%). The following schematic diagram (Figure 14) shows the fixed bed column reactor system [116].

Bioremediation of Nitroaromatic Compounds http://dx.doi.org/10.5772/61253

Figure 14. Schematic representation of the reactor used for the degradation of 4-nitrobenzoate and 4-aminobenzoate by B. cepacia PB4 in the abovementioned study [59].

A similar kind of experiment was performed in year 1999 in which B. cepacia strain PB4 was isolated from 4-aminobenzoate. This also possessed the capability to degrade 4-nitrobenzoate. Thus, considering both classes of the contamination to be toxic and mutagenic, an efficient strategy of decontamination was applied because B. cepacia was able to use both as the sole source of carbon, nitrogen, and energy. In order to prevent toxic effects, these compounds were supplied in lower concentration, i.e., 10–100 ppm and also in order to increase the efficiency of the procedure at such low concentration the degradative bacteria was immobilized on porous diatomaceous celite. This degradation was carried out in a packed bed reactor (PBR). The bioreactor consisted of a glass tube (296 × 41 mm) filled with 150 mm packed bed of celitegrade R-633 or R-635, which was placed at 30°C. It was shown that the nitroaromatic and aromatic amino compounds, which are otherwise unlikely to degrade together if present in any affected area, had been degraded simultaneously by single microorganism supply. The eventual objective of all the biochemical and molecular characterization of the bacterial degradation of pollutants is to develop strains, which could be used in the bio remediation process. In this respect, another good field trial experiment was described by Labana et al. [117, 118] with bacterial strain Arthrobacter protophormiae RKJ100. The result clearly showed that the disappearance of the nitroaromatic pollutants. Similarly, Pseudomonas sp. ST53 was also used as a microbe to degrade TNT and other explosives, but it is best suited on land and water only when the contamination is low [119]. Qureshi et al. [120] reported a bacterial strain Arthrobacter sp. HPC1223, which was capable of degrading 2,4,6-trinitrophenol prominently. This also poses the capability to degrade dinitro‐ phenol and mononitrophenol showing broad substrate specificity.

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3. Evolution of genes for nitroaromatic degradation Nitroaromatics are recent compounds present in the environment and bacterial strains that adapted themselves for the removal of these compounds. This was possible only through evolution of its degradation system at a genetic level. 3.1. Modes of gene evolution 3.1.1. Mutational drift Substrate profile of an existing enzyme may be altered by point mutations in its corresponding genes [124]. The reasons for changes in primary DNA sequences are slippage of DNA poly‐ merase while replications, erroneous DNA repair, and gene conversion [122]. However, results of these changes are relatively smaller. These alone cannot be accountable for adaptation to the new environment by bacteria [123, 121]. 3.1.2. Genetic rearrangement within a cell The rearrangement of genes for the development of new pathway may take place by the help of cells own recombination system. Gene segments can be exchanged between two positions flanked by homologous sequences, insertion elements, transposons, and even sequence identities of four base pairs are sufficient to facilitate this process [122]. 3.1.3. Horizontal gene transfer Horizontal gene transfer is reported as the main source of evolution of pathways in bacteria [123, 124]. Sequencing results of genomes from different bacterial strains have revealed the presence of acquired genes in mosaic like fashion throughout bacterial genomes. Their presence varies from almost negligible (Rickettsia prowazekii and Mycobacterium genitalium) to about 17% (in Synechocystis strain) [123]. Plasmids also play a major role in carrying catabolic genes during horizontal gene transfer [125]. Transposons are also known to facilitate the catabolic gene transfer processes [126]. 3.2. Nitroaromatic degradation pathway as a role model for study the gene evolution A role model to study evolution of microbial pathways is to study the degradation of nitroar‐ omatic compounds in different bacteria. Nitroaromatic compounds are relatively new to the environment, but bacterial systems have already evolved the ability to metabolize them. This cannot be possible only by spontaneous, independent evolution of several new enzymes in a single bacterium. Horizontal gene transfer has to play a key role in combination with the mutagenesis of the existing enzymes to facilitate rapid evolution of new pathways. Evolution of diverse pathways for the degradation of different nitroaromatics thus stands testament to this.

Bioremediation of Nitroaromatic Compounds http://dx.doi.org/10.5772/61253

A good example of this is the evolution of chloronitrobenzene dioxygenase system from a chloronitrobenzene degrading strain Pseudomonas stutzeri [127]. It has several insertion sequences embedded between the gene clusters, which proves such involvement in its evolution. In its dioxygenase enzyme system, reductase and ferredoxin seems to have come from different origins because its reductase and ferredoxin share maximum identity with anthranilate dioxygenase, which is a type IV oxygenase, whereas its oxygenase subunits show a maximum identity with nag (naphthalene degradation gene) and nitroarene dioxygenase, which falls under type III oxygenase systems. This enzyme system thus best illustrates the evolution of catabolic genes best because its upper pathway enzymes seem to have originated from a nitroarene degradation pathway and its lower pathway genes have evolved from some chloroaromatic compound degradation pathway. The genes responsible for these pathways are present in a patchwork like assembly in P. stutzeri [127]. The presence of insertion elements in the gene cluster confirms its role in the formation of a modular assembly and its role in evolution of the gene cluster. Another example is the origin of 2,4-DNT degradation pathway in B. cepacia R34 where enzymes for the degradation pathway have originated from at least three different sources [128]. The first enzyme in the degradation pathway (DNTDO), which removes the first nitrite from 2,4-DNT, seems to have originated from naphthalene degradation pathway like in Ralstonia sp. strain U2. [129]. However, 4-methyl-5-nitrocatechol mono-oxygenase, which facilitates the removal of the second nitrite, appears to be derived from a pathway for degra‐ dation for chloroaromatic compounds [128]. The last enzyme of the pathway could have originated from a gene cluster for amino acid degradation. This is known as progressive compaction of the genes. However, the presence of ORFs without any known role in 2,4-DNT degradation and truncated transposons in the regions suggests that compaction is probably in an intermediate stage in the evolution of such an optimal system with the genetic materials from different bacterial origin. A good example of similar type of evolution is reported in diphenylamine degrader Burkholde‐ ria sp. strain JS665 [130]. In this pathway, diphenylamine is converted to catechol and aniline. An analysis of this sequence of diphenylamine degrading enzyme system showed that it has evolved by recruiting two pathway enzymes, one of which is from dioxygenase and the other is from nitroaniline degradation pathway enzymes in a much more recent evolution event [127]. Another example of evolution of genes for nitrotoluenes degradation is the evolution of 3NTDO in Diaphorobacter sp. strain DS2. Five complete ORFs were identified by probable ORF finding program and by homology to polypeptide sequences from several previously reported multicomponent dioxygenase systems. The predicted translation products from ORFs were designated as a putative regulatory protein, a ferredoxin reductase subunit, a ferredoxin subunit, an oxygenase large, and a small subunit based on their homology. A 571-bp DNA stretch was present in between reductase and ferredoxin subunits. In its gene structure, the regulatory protein is divergently transcribed from the other four ORFs. The organization of gene cluster and its similarity with other known dioxygenases is shown in Figure 15. It has been suggested in several reports that Ralstonia sp. strain U2 [129] is the progenitor of all the nitroarene dioxygenases because it has the entire functional gene in its gene assembly.

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This seems true for the 3NTDO as well because the sequence present in between the reductase and the ferredoxin is truncated into parts of two functional ORFs present in the Ralstonia sp. strain U2. The regulatory protein sequence of 3NTDO (MntR) differs only at three amino acid positions from NagR of Ralstonia sp. strain U2 out of which two are uniquely present in strain DS2 only.

Figure 15. Comparison of gene arrangements in various multicomponent dioxygenases.

The different components (reductase, ferredoxin, and oxygenase) of 3NTDO show different levels of sequence identity with components from similar multicomponent enzyme systems of different organisms. Its reductase subunit (MntAa) shares a high amino acid sequence identity with those of DNTDO from B. cepacia [128] and NDO of Ralstonia sp. strain U2 (99%) [129], but its ferredoxin subunit (MntAb) is 100% identical to the ferredoxin of dinitrotoluene dioxygenase from Burkholderia sp. strain DNT and Burkholderia sp. strain R34. Its large oxygenase subunit (MntAc) showed more identity with chloronitrobenzene dioxygenase (CnbAc, 96%) from P. stutzeri ZWLR2-1, whereas the small oxygenase subunit (MntAd) showed more identity with PahAd of Comamonas testosteroni (96%) and NTDO from Acidovor‐ ax sp. strain JS42. It is known that oxygenase large subunit controls substrate specificity. If we compare important active site residues in oxygenase large subunit (MntAc) of strain DS2 with well-characterized oxygenase systems, it contains amino acid combinations of other systems, in which the sequence retains His293, which is present in NDO system of P. putida 9816-4, Ralstonia U2, C. testosteroni H, and Burkholderia sp. C3. Position 350 is occupied by Valine, which

Bioremediation of Nitroaromatic Compounds http://dx.doi.org/10.5772/61253

is reported in the DNTDO of Burkholderia sp. strain R34. Thus, the above facts seem to indicate that 3NTDO gene in Diaphorobacter sp. strain DS2 came through a horizontal gene transfer from ancestors common to strains like Ralstonia U2 or Burkholderia sp. strain R34, and then its catalytic subunit has been diversely evolved to degrade other nitroaromatic compounds. The mechanism by which enzymes for the degradation of synthetic compounds have evolved so rapidly still cannot be explained only by horizontal gene transfer and mutations. It can be explained in part by the term promiscuity. Promiscuity refers to the ability of a protein to perform dual functions using same active site [131, 132]. Protein evolution toward a new function based on promiscuity involves transition of an existing specialized enzyme to a generalized intermediate enzyme and then into a new specialized enzyme (Figure 16). A good example of this is transcriptional regulator found in nitroarene dioxygenases [133].

Figure 16. Evolution of new specialized enzymes from existing one. (Modified from Ju et al. [134])

The correct functioning of a pathway depends not only on having enzymes with appropriate catalytic activity but also on regulators, which control the expression of the catabolic genes in response to the compounds to be degraded. For example, ntdR, which controls 2NTDO expression in Acidovorax JS42, is supposed to have evolved from an ancestral nagR regulator of naphthalene degradation pathway present in Ralstonia U2. NtdR differs only by five amino acids with nagR. Ju et al. [134] showed how ntdR like regulator could be created from nagR by making mutations at each of the five positions separately and in combinations in a stepwise manner. They also showed that each mutation broadened the effectors range in a stepwise manner without losing the original activity. Both NagR and NtdR can activate transcription in the presence of salicylate, which is a natural inducer of naphthalene degradation genes in strain Ralstonia sp. U2, but ntdR could have gained a broader effector specificity to recognize several nitroaromatic compounds too [134]. Hence, the evolution of the regulatory system of the 2NTDO is in an intermediate stage because it can be induced in response to several nonmetabolizable compounds. Thus, the selection of ntdR variant with high specificity for 2NT with loss of specificity for salicylate would enable the identification of mutations that can lead to the specialized transcription factor from an intermediate stage. Another regulator was reported by Singh et al. [83], where the regulatory protein sequence of 3NTDO (MntR) differs only at three amino acid positions from NagR of Ralstonia sp. strain U2, out of which two are uniquely present in strain DS2 only.

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It can be concluded that the gene evolution in these dioxygenase systems cannot be explained by considering only one mode of evolution. All the modes of evolution (like horizontal gene transfer, selective mutation, and promiscuity) are responsible for the evolution of a dioxyge‐ nase system [133, 135]. Further, the presence of truncated ORFs (which is not required for enzyme activity) reveals that gene evolution is in an intermediate stage of the so-called progressive compaction of the genes.

Acknowledgements We thank IIT Kanpur and our collaborators for their help during this work on Diaphorobacter isolation and characterization. KM thanks CSIR for a senior research fellowship.

Author details Deepak Singh, Keerti Mishra and Gurunath Ramanthan* *Address all correspondence to: [email protected] Department of Chemistry, Indian Institute of Technology Kanpur, Kanpur, India

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[56] Razo-Flores E. et al. Biotransformation and biodegradation of N-substituted aromat‐ ics in methanogenic granular sludge. FEMS Microbiol Review. 1997;20:525–538. [57] McCormick N, Feeherry F E, Levinson H S. Microbial transformation of 2,4,6-trinitro‐ toluene and other nitroaromatic compounds. Appl Environ Microbiol. 1976;31:949– 958. [58] Donlon B A, Razo-Flores E, Lettinga G, Field J A. Continuous detoxification, transfor‐ mation and degradation of nitrophenols in upflow anaerobic sludge blanket (USAB) reactors. Biotechnol Bioeng. 1996;51:439–449. [59] Peres C M. et al. Continuous degradation of mixtures of 4-nitrobenzoate and 4-ami‐ nobenzoate by immobilized cells of Burkholderia cepacia strain PB4. Appl Microbiol Biotechnol. 1999;52:440–445. [60] Rieger P G, Knackmuss H J. Basic knowledge and perspectives on biodegradation of 2,4,6 trinitrotoluene and related nitroaromatic compounds in contaminated soil. In: J.C. Spain (ed.), Biodegradation of Nitroaromatic Compounds. New York: Plenum Press, 1995,1–18. [61] Nishino S F, Spain J C, He Z. Strategies for aerobic degradation of nitroaromatic com‐ pounds by bacteria: process discovery to field application. Spain J C, Hughes J B, Knackmuss H J (eds.), Biodegradation of Nitroaromatic Compounds and Explosives. New York: Lewis Publ; 2000,7–61. [62] Simpson J R, Evans W C. The metabolism of nitrophenols by certain bacteria. Bio‐ chem J. 1953;55:XXIV. [63] Ecker S. et al. Catabolism of 2,6-dinitrophenol by Alcaligenes eutrophus JMT134 and 222. Arch Microbiol, 1992;158:149–154. [64] Haigler B E, Wallace W H, Spain J C. Biodegradation of 2-nitrotoluene by Pseudomo‐ nas sp. strain JS42. Appl Environ Microbiol. 1994,60,3466–3469. [65] Mulla S I. et al. Biodegradation of 2-nitrotoluene by Micrococcus sp.strain SMN-1. Bio‐ degradation. 2011;22:95–102. [66] Singh D, Ramanathan G. Biomineralization of 3-nitrotoluene by Diaphorobacter spe‐ cies. Biodegradation. 2013;24:645–655. [67] Nishino S F, Spain J C. Oxidative pathway for the biodegradation of nitrobenzene by Comamonas sp. strain JS765. Appl Environ Microbiol. 1995;61:2308–2313. [68] Suen W C, Spain J C. Cloning and characterization of Pseudomonas sp. strain DNT genes for 2,4-dinitrotoluene degradation. J Bacteriol. 1993;175:1831–1837. [69] Groenewegen P E J, De Bont J A M. Degradation of 4-nitrobenzoate via 4-hydroxyla‐ minobenzoate and 3,4-dihydroxybenzoate in Comamonas acidovorans NBA-10. Arch Microbiol. 1992;158:381–386.

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[70] Ali-Sadat S, Mohan K S, Walia S K. A novel pathway for the biodegradation of 3-ni‐ trotoluene in Pseudomonas putida. FEMS Microbiol Ecol. 1995;17:169–176. [71] Williams R W, Taylor S C, Williams P A. A novel pathway for the catabolism of 4nitrotoluene by Pseudomonas. J Gen Microbiol, 1993;139:1967–1972. [72] Spiess T. et al. A new 4-nitrotoluene degradation pathway in a Mycobacterium strain. Appl Environ Microbiol. 1998;64:446. [73] An D, Gibson D T, Spain J C. Oxidative release of nitrite from 2-nitrotoluene by a three-component enzyme system from Pseudomonas sp.strain JS42. J Bacteriol. 1994;176:7462–7467. [74] Jain R K, Dreisbach J H, Spain J C. Biodegradation of p-nitrophenol via 1,2,4-benze‐ netriol by Arthrobacter sp. Appl Environ Microbial. 1994;60:3030–3032. [75] Spain J C, Gibson D T. Pathway for biodegradation of p-nitrophenol in a Moraxella sp. Appl Environ Microbiol. 1991;57:812–819. [76] Zeyer J, Kearney P C. Degradation of o-nitrophenol and m-nitrophenol by a Pseudo‐ monas putida. J Agric Food Chem. 1984;32:238–242. [77] Delgado A. et al. Nitroaromatics are substrates for the TOL plasmid upper-pathway enzymes. Appl Environ Microbiol. 1992;58:415–417. [78] Robertson J B. et al. Oxidation of nitrotoluenes by toluene dioxygenase: evidence for a monooxygenase reaction. Appl Environ Microbiol. 1992;58:2643–2648. [79] Struijs J, Stoltenkamp J. Ultimate biodegradation of 2-, 3-, and 4-nitrotoluene. Sci To‐ tal Environ. 1986;57:161–170. [80] Goodall J L, Peretti S W. Dynamic modeling of meta- and para-nitrobenzoate metab‐ olism by a mixed co-immobilized culture of Comamonas sp. JS46 and JS47. Biotechnol Bioeng, 1998;59:507–516. [81] Williams R W, Taylor S C, Williams P A. A novel pathway for the catabolism of 4nitrotoluene by Pseudomonas. J Gen Microbiol, 1993;139:1967–1972. [82] Walia S K, Ali-Sadat S, Rasul Chaudhry G. Influence of nitro group on biotransfor‐ mation of nitrotoluenes in Pseudomonas putida strain OU83.Pestic Biochem Physiol. 2003;76: 73–81. [83] Singh D, Kumari A, Ramanathan G. 3-Nitrotoluene dioxygenase from Diaphorobacter sp. strains: cloning, sequencing and evolutionary studies. Biodegradation. 2014;25:479–492. [84] Singh, D. et al. Expression, purification and substrate specificities of 3-nitrotoluene dioxygenase from Diaphorobacter sp. strain DS2. Biochem Biophys Res Commun. 2014;445:36–42.

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[85] Spanggord R J. et al. Biodegradation of 2,4-dinitrotoluene by a Pseudomonas sp. Appl Environ Microbiol. 1991;57:3200–3205. [86] Nishino S F, Paoli G C, Spain J C. Aerobic degradation of dinitrotoluenes and path‐ way for bacterial degradation of 2,6-dinitrotoluene. Appl Environ Microbiol. 2000;66:2139–2147. [87] Nishino S F, Spain J C. Technology status review: bioremediation of dinitrotoluene (DNT), 2001 [88] Lendenmann U, Spain J C. Simultaneous biodegradation of 2,4-dinitrotoluene and 2,6-dinitrotoluene in an aerobic fluidizedbed biofilm reactor. Environ Sci Technol. 1998;32: 82–87. [89] Bae B, Autenrieth R L, Bonner J S. Aerobic biotransformation and mineralization of 2,4,6-trinitrotoluene. In: R.E. Hinchee, R.E. Hoeppel, D.B. Anderson (eds.), Bioreme‐ diation of Recalcitrant Organics. Columbus, Ohio: Battelle Press; 1995,231–238. [90] Paca J,Halecky M, Barta J, Bajpai R. Aerobic biodegradation of 2,4-DNT and 2,6DNT: Performance characteristics and biofilm composition changes in continuous packed-bed bioreactors. J Hazard Mater. 2009;163:848–854. [91] Wang Z Y, Ye Z F, Zhang M H. Bioremediation of 2,4-dinitrotoluene (2,4-DNT) in im‐ mobilized micro-organism biological filter. J Appl Microbiol. 2011;110:1476–1484. [92] Lenke H, Achtnich C, Knackmuss H J. Perceptives of bioelimination of polynitroaro‐ matic compounds. Spain J C, Highes J B, Knackmuss H J (eds.), Biodegradation of Ni‐ troaromatic Compounds and Explosives. Boca Raton, FL: Lewis Publishers. 2000, 91– 126. [93] Huang S. et al. 2,4,6-Trinitrotoluene reduction by carbon monoxide dehydrogenase from Clostridium thermoaceticum. Appl Environ Microbiol. 2000;66:1474–1478. [94] Van Aken B. et al. Transformation and mineralization of 2,4,6-trinitrotoluene (TNT) by manganese peroxidase from the white-rot basidiomycete Phlebia radiata. Biodegra‐ dation. 1999;10:83–91. [95] French C E, Nicklin S, Bruce N C. Aerobic degradation of 2,4,6-trinitrotoluene by En‐ terobacter cloacae PB2 and by pentaerythritol tetranitrate reductase. Appl Environ Mi‐ crobiol. 1998;64:2864–2868. [96] Haidour A, Juan L R. Identification of products resulting from the biological reduc‐ tion of 2,4,6-trinitrotoluene, 2,4-dinitrotoluene, and 2,6-dinitrotoluene by Pseudomo‐ nas sp. Environ Sci Technol. 1996;30:2365–2370. [97] Shim C Y, Crawford D L. Biodegradation of trinitrotoluene (TNT) by a strain of Clos‐ tridium bifermentans. Hinchee R E, Fredrickson J, Alleman B C (eds.), Bioaugmenta‐ tion for Site Remediation. Columbus, Ohio: Battelle Press; 1995,57–69.

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[98] Lewis T A. et al. Microbial transformation of 2,4,6-trinitrotoluene. J Ind Microbiol Bi‐ otechnol. 1997;18:89–96. [99] Boopathy R, Kulpa C F. Trinitrotoluene as a sole nitrogen source for a sulfate-reduc‐ ing bacterium Desulfovibrio sp. (B strain) isolated from an anaerobic digester. Curr Microbiol. 1992;25:235–241. [100] Preuss A, Fimpel J, Dickert G. Anaerobic transformation of 2,4,6-trinitrotoluene (TNT). Arch Microbiol. 1993;159:345–353. [101] Drzyzga O. et al. Mass balance studies with 14C-labeled 2,4,6-trinitrotoluene (TNT) mediated by an anerobic Desulfovibrio species and an aerobic Serratia species. Curr Microbiol. 1998;37:380–386. [102] Boopathy R, Kulpa C F. Biotransformation of 2,4,6-trinitrotoluene (TNT) by a Metha‐ nococcus sp. (strain B) isolated from a lake sediment. Can J Microbiol. 1994;40: 273– 278. [103] Esteve-Núñez A, Ramos J L. Metabolism of 2,4,6-trinitrotoluene by Pseudomonas sp. JLR11. Environ Sci Technol. 1998;32:3802–3808. [104] Kalafut T. et al. Biotransformation patterns of 2,4,6-trinitrotoluene by aerobic bacte‐ ria. Curr Microbiol. 1998;36:45–54. [105] French C E, Nicklin S, Bruce N C. Aerobic degradation of 2,4,6-trinitrotoluene by En‐ terobacter cloacae PB2 and by pentaerythritol tetranitrate reductase. Appl Environ Mi‐ crobiol. 1998;64:2864–2868. [106] Alvarez M A. et al. Pseudomonas aeruginosa strain MA01 aerobically metabolizes the aminodinitrotoluenes produced by 2,4,6-trinitrotoluene nitro group reduction. Can J Microbiol. 1995;41:984–991. [107] Pak J W. et al. Transformation of 2,4,6-trinitrotoluene by purified xenobiotic reduc‐ tase B from Pseudomonas fluorescens I-C. Appl Environ Microbiol. 2000;66:4742–4750. [108] Naumova R P, Selivanovskaya S L U, Mingatina F A. Possibilities for the deep bacte‐ rial destruction of 2,4,6-trinitrotoluene. Mikrobiologia. 1988;57:218–222. [109] Gilcrease C P, Murphy V G. Bioconversion of 2,4-diamino-6-nitrotoluene to a novel metabolite under anoxic and aerobic conditions. Appl Environ Microbiol. 1995;61:4209–4214. [110] Fiorella P D, Spain J C. Transformation of 2,4,6-trinitrotoluene by Pseudomonas pseu‐ doalcaligenes JS52. Appl Environ Microbiol. 1997;63:2007–2015. [111] Vorbeck C. et al. Initial reductive reactions in aerobic microbial metabolism of 2,4,6trinitrotoluene. Appl Environ Microbiol. 1998;64:246–252. [112] Montpas S. et al. Degradation of 2,4,6-trinitrotoluene by Serratia marcescens. Biotech‐ nol Lett. 1997;19:291–294.

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[113] Collie S. et al. Degradation of 2,4,6-trinitrotoluene (TNT) in an aerobic reactor. Che‐ mosphere. 1995;31:3025–3032. [114] Cho Y S, Lee B U, Oh K H. Simultaneous degradation of nitroaromatic compounds TNT, RDX, atrazine, and simazine by Pseudomonas putida HK-6 in bench-scale bio‐ reactors. J Chem Technol Biotechnol. 2008;83:1211–1217. [115] Oh K, Kim Y. Degradation of explosive 2,4,6-trinitrotoluene by s-triazine degrading bacterium isolated from contaminated soil. Bull Environ Contam Toxicol, 1998;61:702–708. [116] Gisi D, Stucki G, Hanselmann K W. Biodegradation of the pesticide 4,6-dinitro-orthocresol by microorganisms in batch cultures and in fixed-bed column reactors. Appl Microbiol Biotechnol. 1997;48:441–448. [117] Labana S. et al. Pot and field studies on bioremediation of p-nitrophenol contaminat‐ ed soil using Arthrobacter protophormiae RKJ100. Environ Sci Technol. 2005;39:3330– 3337. [118] Labana S. et al. A microcosm study on bioremediation of p-nitrophenol-contaminat‐ ed soil using Arthrobacter protophormiae RKJ100. Appl Microbiol Biotechnol. 2005;68:417–424. [119] Snellinx Z. et al. Biological remediation of explosives and related nitroaromatic com‐ pounds. Environ Sci Pollut Res Int. 2002;9:48–61. [120] Qureshi A, Kapley A, Purohit H J. Degradation of 2,4,6-trinitrophenol (TNP) by Ar‐ throbacter sp. HPC1223 isolated from effluent treatment plant. Indian J Microbiol. 2012;52:642–647. [121] Van der Meer J R. et al. Molecular mechanisms of genetic adaptation to xenobiotic compounds. Microbiol Rev. 1992;56:677–694. [122] van der Meer J R. Evolution of novel metabolic pathways for the degradation of chloroaromatic compounds.Antonie Van Leeuwenhoek, 1997;71:159–178. [123] Ochman H, LawrenceJ G, Groisman E A. Lateral gene transfer and the nature of bac‐ terial innovation. Nature. 2000;405:299–304. [124] Koonin E V, Makarova K S, Aravind L. Horizontal gene transfer in prokaryotes: quantification and classification. Ann Rev Microbiol. 2001;55:709–742. [125] Alexander M. Biodegradation and Bioremediation. 2nd ed. San Diego, CA: Academic Press; 1999. [126] Tan H M. Bacterial catabolic plasmids. Appl Microbiol Biotechnol. 1999;51:1–12. [127] Liu H. et al. Patchwork assembly of nag-like nitroarene dioxygenase genes and the 3chlorocatechol degradation cluster for evolution of the 2-chloronitrobenzene catabo‐

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lism pathway in Pseudomonas stutzeri ZWLR2-1. Appl Environ Microbiol. 2011;77:4547–4552. [128] Johnson G R, Jain R K, Spain J C. Origins of the 2,4-dinitrotoluene pathway. J Bacter‐ iol. 2002;184:4219–4232. [129] Fuenmayor S L. et al. A gene cluster encoding steps in the conversion of naphthalene to gentisate in Pseudomonas sp. strain U2. J Bacteriol. 1998;180:2522–2530. [130] Shin K A, Spain J C. Pathway and evolutionary implications of diphenylamine biode‐ gradation by Burkholderia sp. strain JS667. Appl Environ Microbiol. 2009;75:2694– 2704. [131] Copley S D. Enzymes with extra talents: moonlighting functions and catalytic pro‐ miscuity. Curr Opin Chem Biol. 2003;7:265–272. [132] James L C, Tawfik D S. Conformational diversity and protein evolution- a 60-yearold hypothesis revisited. Trends Biochem Sci. 2003;28:361–368. [133] Kivisaar M. Degaradation of nitroaromatic compounds: a model to study evolution of metabolic pathways. Mol Microbiol. 2009;74:777–781. [134] Ju K S, Parales J V, Parales R E. Reconstructing the evolutionary history of nitroto‐ luene detection in the transcriptional regulator NtdR. Mol Microbiol. 2009;74:826– 843. [135] Kivisaar, M. Evolution of catabolic pathways and their regulatory systems in syn‐ thetic nitroaromatic compounds degrading bacteria. Mol Microbiol. 2011;82:265–268.

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Chapter 3

Selection of Promising Bacterial Strains as Potential Tools for the Bioremediation of Olive Mill Wastewater Daniela Campaniello, Antonio Bevilacqua, Milena Sinigaglia and Maria Rosaria Corbo Additional information is available at the end of the chapter http://dx.doi.org/10.5772/60896

Abstract The main objective of this paper was the selection of promising bacterial strains to be used as potential tools to remove phenols in olive mill wastewater (OMW) or in other food wastes. Therefore, 12 OMW samples were analyzed and 119 isolates were collected. After a preliminary screening on a medium containing vanillic and cinnamic acids, three isolates were selected to evaluate their viability in presence of different compounds (cinnamic, vanillic and caffeic acids, rutin, tyrosol and oleuropein) and a possible bioremediation effect. The isolates generally survived with phenols added and exerted a significant bioremediation activity in some samples (reduction of phenols by 20%). The last step was focused on the evaluation of the combined effects of pH, cinnamic and vanillic acids on the viability of a selected isolate (13M); the combination of the acids exerted a strong effect on the target, but alkaline pH played a protective role. Keywords: Bioremediation, phenol degradation, phenolic compounds, olive mill wastewater

1. Introduction Olive oil production is one of the most important food sectors in the Mediterranean area as olive processing is considered a traditional industry for its countries since ancient times [1]. It is mainly produced in Spain (36% of the global production), Italy (24%), Greece (17%) followed

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by Portugal, France, Cyprus, Croatia, Turkey, Syria, and Tunisia. New producers are Argen‐ tina, Australia, and South Africa. Olives are processed through two methods: pressing (discontinuous process) and centrifuging (continuous process, two/three phase centrifugation). The main inconvenience of these methods is the production of a polluting by-product that is a dark effluent known as olive mill wastewater (OMW) [2]. The environmental impact of olive oil production is strong due to the use of large quantities of water and the production of OMW: 1,000 kg of olives produce 0.5 m3–1.5 m3 of OMWs [1]. They are by-products generally considered undesirable but inevitable for every olive processing. As defined in reference [3], OMW is “a stable emulsion constituted by vegetation waters of the olives, water from the processing, olive pulp and oil.” It is characterized by a particular color (intensive violet-dark brown up to black color), odor (strong specific olive oil smell), high degree of organic pollution (expressed as biological and chemical oxygen demand (BOD and COD) values), acidic pH, high electrical conductivity, high content of polyphenols, high buffer capacity, and high content of solid matter. OMWs are generally composed of water (83%–96%) and organic fraction (3.5%–15%) com‐ posed of 1%–8% carbohydrates, 0.5%–2.4% nitrogen compounds, 0.5%–1.5% organic acids, 0.02%–1% of fatty acids such as propionic, butyric, etc., and 1%–1.5% of phenolic compounds consisting of a hydroxyl group (-OH) bound directly to an aromatic hydrocarbon group and pectins [4]. Concerning phenols, they comprise low molecular weight compounds and polyphenols. Low molecular weight compounds are represented by caffeic, cinnamic, 2,6-dihydroxybenzoic, phydroxybenzoic, syringic, 3,4,5-trimethoxybenzoic, vanillic, and veratric acids; they have phytotoxic effects and antibacterial activity. Polymeric phenols (lignins, tannins, etc.) cause the typical brownish-black color of OMW and are the most recalcitrant fraction of this effluent. The quantitative and qualitative composition of OMWs are variable due to climatic conditions, variety, ripeness of olives, and extraction processes; generally, they are produced in high quantities in a short time, thus their disposal represents an important problem. As OMWs are rich in nutrients they could be used to remediate arid or semi-arid regions but their phytotoxicy affect plant growth [5]. OMWs have the highest polluting rate within the food industry due to the fact that they are recalcitrant to traditional biodegradative methods. The reduction of COD and BOD values represent an important goal for many industries but the high content in phenols complicate waste management; they exert an antimicrobial activity towards wastewater microflora thus biodegradation is slowed [6]. For these reasons phenols are considered as undesirable compounds; thus, physical, chemical, or biological treatments are used to reduce their pollutant load. Waste remediation has been traditionally performed through some expensive methods (incineration, pyrolysis, landfill, etc.). In recent years, the increasing trends towards green economy and friendly approaches for the environment are the background to design alterna‐

Selection of Promising Bacterial Strains as Potential Tools for the Bioremediation of Olive Mill Wastewater http://dx.doi.org/10.5772/60896

tive ways. According to this point of view, numerous researchers proposed bioremediation, defined as “the process whereby organic wastes are biologically degraded under controlled conditions to an innocuous state or to level below the limits established by regulatory author‐ ities” [7]. According to the Environmental Protection Agency (EPA), bioremediation is a “treatment that use naturally occurring organisms to break down hazardous substances into less toxic or nontoxic substances.” Thanks to their ubiquity and metabolic pathways (aerobic, anaerobic fermentation, and cometabolism) microorganisms are able to degrade and utilize various toxic compounds as energy source. Generally, the aerobic biodegradation has a higher efficiency than anaerobic processes and it is widely used. Nevertheless, in many cases, aerobic and anaerobic processes can also be used in series to reduce the complexity or the toxicity of the contaminants. Numerous bacteria such as Bacillus pumilus [8], Pediococcus pentosaceus [9], Lactobacillus plantarum [10], Arthrobacter sp. [11], Azotobacter vinelandii [12–14], Azotobacter chroococcum [15], Pseudomo‐ nas putida, and Ralstonia sp. [16, 17] were able to degrade and/or remove phenols from OMW. Yeasts and molds are also able to degrade phenols, namely, Candida tropicalis, Candida cylin‐ dracea and Yarrovia lypolitica [18, 19, 3, 20, 4], and white-rot fungi such as Phanerochaete chrysosporium or the genus Pleurotus [21–25]. In addition, Trametes versicolor, Funalia trogii, Lentinus edodes, Aspergillus niger, and Aspergillus terreus have been also mentioned as phenoldegrading organisms [26]. The main objective of this paper was the selection of promising bacterial strains to be used as potential tools for bioremediation; namely, after the isolation of some strains from OMWs, they were studied in relation to their ability to grow in a medium containing two secondary phenols. Then, a validation on a lab scale was performed.

2. Materials and methods 2.1. Isolation and phenotyping of potential phenol-degrading strains Twelve different samples of OMW were analyzed. Aliquots of 100 ml of each OMW sample were mixed with 900 ml of sterile Ringer solution (0.25×; Oxoid, Milan, Italy) and shaken at 100 rpm for 30 min at room temperature. Then, this homogenate was serially diluted with a sterile saline solution (0.9% NaCl) and plated onto a Plate Count Agar (PCA; Oxoid) at 30°C for 48–72 h (for mesophilic bacteria and Bacillus) and Pseudomonas Agar Base + Pseudomonas C-F-C supplement (Oxoid) at 25°C for 72 h for Pseudomonadaceae. The analyses were performed in duplicate over two different batches. From each batch, some colonies were randomly selected from plates, and stored at 4°C on Tryptone Soya Agar (TSA) slants (Oxoid, Milan, Italy). Phenotyping of isolates was carried out through different tests (Gram, catalase activity, oxidase test, proteolytic activity, and oxido-fermentation). 2.2. Selection of potential phenol-degrading strains The ability of the isolates to grow with phenols added was evaluated using Mineral Salt Medium (MSM), a synthetic medium containing K2HPO4 (1.6 g/l), KH 2PO4 (0.4 g/l), NH 4NO3

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(0.5 g/l), MgSO4*7H2O (0.2 g/l), CaCl2 (0.025 g/l), FeCl2 (0.005 g/l), Agar (12 g/l), cinnamic (C9H8O2) or vanillic acids (C8H8O4) (0.5 or 1 g/l; Sigma-Aldrich, Milan, Italy). After streaking the isolates onto the surface of this modified medium, the plates were incubated at 30°C for mesophilic bacteria and Bacillus and at 25°C for Pseudomonadaceae for 72 h. MSM without phenolic compounds was used as control. For a second assay, the isolates were preliminary grown in MSM broth added of 0.025 g/l and 0.05 g/l of cinnamic and vanillic acids and incubated for 24 h; thereafter they were streaked onto the surface of MSM with phenols, as reported above. 2.3. Effect of phenolic compounds on the viable cell count This step was performed on three selected strains (6P, 13M, and 44M); they were grown in TS broth incubated at 25°C (strain 6P) or at 30°C (strains 13M and 44M) for 48 h. Each culture was centrifuged at 4,000 rpm for 10 min and the pellet was re-suspended in sterile saline solution (0.9% NaCl); ca. 6–7 log cfu/ml were inoculated in MSM medium (1, 2, and 3 g/l), added with phenols (cinnamic acid; vanillic acid; caffeic acid-C9H9O4; rutin hydrate-C27H30O16 x H2O; tyrosol-C8H10O2; oleuropein-C25H32O13; phenols were purchased from Sigma-Aldrich). MSM without phenols was used as control. The samples were stored at 25°C–30°C for 33 days and periodically analyzed to evaluate the viable count on TSA and the content of phenols through the Folin-Ciocalteau method [27]. The analyses were performed in duplicate and the results analyzed through one-way Analysis of Variance (one-way ANOVA), using Tukey’s test as the post-hoc comparison test, or t-student test for paired comparisons. The statistical analysis was performed using the software Statistica for Windows version 10.0 (Statsoft, Tulsa, OK, USA). 2.4. Combined effects of pH, cinnamic, and vanillic acids on the viable count of the strain 13M The strain 13M was used as target; it was inoculated into MSM broth to 6–7 log cfu/ml. The amounts of cinnamic and vanillic acids and pH varied according to a 2k experimental (Table 1). The samples were stored at 30°C and periodically analyzed to evaluate the viable count and phenol content (up to 12 days). Combinations

pH

Cinnamic acid (g/l)

Vanillic acid (g/l)

A

7

0

0

B

7

0

2

C

7

2

0

D

7

2

2

E

9

0

0

F

9

0

2

G

9

2

0

H

9

2

2

Table 1. Combinations of 2k design.

Selection of Promising Bacterial Strains as Potential Tools for the Bioremediation of Olive Mill Wastewater http://dx.doi.org/10.5772/60896

The analyses were performed in duplicate and the results of viable count analyzed through a multiple regression approach by using the option DoE/2k design of the software Statistica for Windows.

3. Results 3.1. Isolation and screening on MSM The viable count of mesophilic and spore-forming bacteria and Pseudomonas was fairly high (7–8 log cfu/ml) (Table 2); thus, we selected 119 isolates (46 labeled as mesophilic bacteria, 44 and 29 belonging to Bacillus and Pseudomonas genera, respectively). Samples 1

2

3

4

5

6

7

8

9

10

11

12

M

7.95

7.77

8.30

7.60

7.48

9.30

8.30

8.30

8.00

6.78

6.30

6.70

B

7.30

6.70

7.30

6.78

6.78

7.00

6.60

6.70

5.70

6.48

6.48

6.30

P

7.70

6.70

8.00

6.90

7.60

6.30

8.30

8.00

7.84

7.00

7.48

8.78

Table 2. Viable count (log cfu/g) of mesophilic bacteria (M), Bacillus (B), and Pseudomonas (P) in OMW samples. Data are the average (n=2).

The isolates were streaked on MSM with cinnamic or vanillic acids; Figure 1 shows the results   for vanillic acid. Namely, 12 isolates were able to grow in presence of 0.5 g/l of this compound Figure 1: Screening of the isolates on MSM with vanillic acid (0.5 g/l and 1.0 g/l). The numbers on  the pictures indicate if the targets are able or not able to grow.  (Figure 1a) and 9 with 1g/l (Figure 1b). None of the isolates grew with cinnamic acid.    

  (a) 

(b)  Figure 2: Screening of the isolates on MSM with vanillic acid (0.5 g/l and 1.0 g/l) after the induction.  Figure 1. Screening of the isolates on MSM with vanillic acid (0.5 g/l and 1.0 g/l). The numbers on the pictures indicate The numbers on the pictures indicate if the targets are able or not able to grow.  if the targets are able or not able to grow.

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The screening was also performed after isolate growing in MSM broth with low amounts of phenols; this step could be referred as an induction phase, aimed at inducing the resistance to phenols. Figure 2 shows the results with 1 g/l of vanillic acid. There were 32 out of 119 isolates that acquired the ability to grow in MSM with vanillic acid; at the lowest concentration (0.5 g/ l) 49 strains were able to grow. The same protocol was also used for cinnamic acid, but only a single isolate was able to grow after induction both at 0.5 g/l and 1 g/l (the isolate 26M).  

  (a) 

  (b)  Figure 3: Viability of the isolate 6P in MSM+caffeic acid (mean values ± standard deviation). *, viable  count in MSM+caffeic acid are significantly different from control.   

Figure 2. Screening of the isolates on MSM with vanillic acid (0.5 g/l and 1.0 g/l) after the induction. The numbers on the pictures indicate if the targets are able or not able to grow.

3.2. Viability and phenol reduction by some selected isolates Three isolates (6P, 13M, and 44M; see Table 3 for their phenotypic traits) were selected and used as targets to assess their viability in the presence of various phenolic compounds (caffeic, cinnamic and vanillic acids, oleuropein, and rutin and tyrosol) at different concentrations (1 g/l, 2 g/l, and 3 g/l); moreover, we focused also on microorganism ability to reduce phenol content. Figure 3 reports the viability of the isolate 6P in the presence of caffeic acid; the initial cell number was 7 log cfu/ml. Then, it underwent a strong reduction within 5 days (ca. 1.5 log cfu/ml at 1 g/l and 2 log cfu/ ml at 3 g/l); in the last days of storage we found a tailing effect, with a residual cell count of 5 log cfu/ml. Similar results were found in the presence of cinnamic acid, tyrosol, rutin, and oleuropein (data not shown). Vanillic acid at 2 g/l reduced the viable count by 3 log cfu/ml in 5 days with a final tailing effect and a residual cell count of 3–4 log cfu/ml. The lowest concentration (1 g/l) resulted in a slower death kinetic, with a similar residual viable count (Figure 4).

Selection of Promising Bacterial Strains as Potential Tools for the Bioremediation of Olive Mill Wastewater http://dx.doi.org/10.5772/60896

Phenotyping

Isolates 6P

13M

44M

Gram

-

+

+

Catalase

+

+

+

Oxydase

-

+

+

Proteolitic activity

-

+

+

O/F

F

F

F

Table 3. Phenotyping of the isolates selected for the second step of the research. F, metabolism under aerobic and anaerobic conditions.

Figure 3. Viability of the isolate 6P in MSM+caffeic acid (mean values ± standard deviation). *, viable count in MSM +caffeic acid are significantly different from control.

Figure 4. Viability of the isolate 6P in MSM+vanillic acid (mean values ± standard deviation). Letters indicate signifi‐ cant differences (one-way ANOVA and Tukey’s test; P200 μm) [17, 19], and this difference could be related to the method used to freeze the gels before lyophilization. In the present study AX gels were frozen by immersion in liquid nitrogen (fast congelation), while previous studies [17, 19] reported AX gel congelation at -20°C for several hours (slow congelation). An important aspect of achieving a high-quality frozen material, particularly with high water content such as gels, is the freezing rate. Fast congelation results in a better-preserved structure (i.e., finer ice crystals). The microstructural characteristics of AX1 and AX2 gels were similar to those reported in AX gels frozen by fast congelation [32, 33] under similar conditions to those used in the present study. In those previous studies [32, 33], AX gels can be also compared with an irregular honeycomb structure in which pore diameters ranged from 12 μm to 50 μm.

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Figure 6. The SEM of lyophilized AX1 (a, b) and AX2 (c, d) gels. a and c at 500x magnification; b and d at 2000x magni‐ fication.

The entrapment of biomolecules or microorganisms in high cell dimension (>200μm) AX gels has been previously reported [14, 19, 31], but reducing cell dimensions of AX gels could increase the possibility of carried out smaller compounds or cells. Therefore, AX1 and AX2 gels microstructural characteristics could be of interest for the development of designed delivery systems, which could allow alternative uses for maize wastewater.

4. Conclusion AX1 and AX2 presented similar molecular identity (FT-IR) and molecular weight distribution, but different FA content. Both AX1 and AX2 presented gelling capability under laccase exposure. The lower FA content in AX2 form gels presenting minor elasticity values and a

Gelation of Arabinoxylans from Maize Wastewater — Effect of Alkaline Hydrolysis Conditions on the Gel... http://dx.doi.org/10.5772/61022

more fragmented microstructure. These results indicate that nixtamalization process condi‐ tions can modify the characteristics of AX gels. Environmental concerns have triggered research on alternative nixtamalization processes rendering residues with less environment impact. AX recovering from this kind of less pollutant maize wastewater could be an inter‐ esting research subject in order to explore the structural and functional properties of this hydrocolloid.

Acknowledgements This research was supported by Fondo de Infraestructura-CONACYT, Mexico (Grant 226082 to E. Carvajal-Millan). The authors are pleased to acknowledge Alma C. Campa-Mada, Karla G. Martínez-Robinson, and Alma R. Toledo Guillén (CIAD) for their technical assistance.

Author details Rita Paz-Samaniego1, Elizabeth Carvajal-Millan1*, Francisco Brown-Bojorquez2, Agustín Rascón-Chu1, Yolanda L. López-Franco1, Norberto Sotelo-Cruz2 and Jaime Lizardi-Mendoza1 *Address all correspondence to: [email protected] 1 Research Center for Food and Development, CIAD, A.C. 83304 Hermosillo, Sonora, Mexico 2 University of Sonora, 83000 Hermosillo, Sonora, Mexico

References [1] Valderrama-Bravo C, Gutiérrez-Cortez E, Contreras-Padilla M, Rojas-Molina I, Mos‐ quera JC, Rojas-Molina A, Beristain F, Rodríguez-García, ME. Constant pressure fil‐ tration of lime water (nejayote) used to cook kernels in maize processing. Journal of Food Engineering. 2012;110(3):478-486. DOI: 10.1016/j.jfoodeng.2011.12.018. [2] Acosta-Estrada BA, Lazo-Vélez MA, Nava-Valdez Y, Gutiérrez-Uribe JA, Serna-Sal‐ dívar SO. Improvement of dietary fiber, ferulic acid and calcium contents in pan bread enriched with nejayote food additive from white maize (Zea mays). Journal of Cereal Science. 2014;60(1):264-269. DOI: 10.1016/j.jcs.2014.04.006. [3] Martínez-Bustos F, Martínez-Flores HE, Sanmartín-Martínez E, Sánchez-Sinencio F, Chang YK, Barrera-Arellano D, Rios E. Effect of the components of maize on the

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quality of masa and tortillas during the traditional nixtamalisation process. Journal of the Science of Food and Agriculture. 2001;81:1455-1462. DOI: 10.1002/jsfa.963. [4] Niño-Medina G, Carvajal-Millán E, Lizardi J, Rascon-Chu A, Marquez-Escalante JA, Gardea A, Martinez-Lopez AL, Guerrero V. Maize processing waste water arabinox‐ ylans: Gelling capability and cross-linking content. Food Chemistry. 2009;115(4): 1286-1290. DOI: 10.1016/j.foodchem.2009.01.046. [5] Gutiérrez-Uribe JA, Rojas-García C, García-Lara S, Serna-Saldivar SO. Phytochemical analysis of wastewater (nejayote) obtained after lime-cooking of different types of maize kernels processed into masa for tortillas. Journal of Cereal Science. 2010;52(3): 410-416. DOI: 10.1016/j.jcs.2010.07.003. [6] Ayala-Soto FE, Serna-Saldívar SO, García-Lara S, Pérez-Carrillo E. Hydroxycinnamic acids, sugar composition and antioxidant capacity of arabinoxylans extracted from different maize fiber sources. Food Hydrocolloids. 2014;35:471-475. DOI: 10.1016/ j.foodhyd.2013.07.004. [7] Martínez-López AL, Carvajal-Millan E, Lizardi-Mendoza J, López-Franco Y, RascónChu A, Salas-Muñoz E, Ramírez-Wong B. Ferulated arabinoxylans as by-product from maize wet-milling process: characterization and gelling capability. In: JimenezLopez JC, editor. Maize cultivation, uses and health benefits. New York: Nova Sci‐ ence Publisher, Inc.; 2012. p. 65–73. [8] Saulnier L, Guillon F, Sado PE, Chateigner-Boutin AL, Rouau X. Plant cell wall poly‐ saccharides in storage organs: Xylans (Food Applications). 2013. DOI: 10.1016/ b978-0-12-409547-2.01493-1. [9] Izydorczyk M, Biliaderis C. Cereal arabinoxylans: advances in structure and physico‐ chemical properties. Carbohydrate Polymers. 1995;28:33-48. [10] Niño-Medina G, Carvajal-Millán E, Rascon-Chu A, Marquez-Escalante JA, Guerrero V, Salas-Muñoz E. Feruloylated arabinoxylans and arabinoxylan gels: structure, sources and applications. Phytochemistry Reviews. 2010;9(1):111-120. DOI: 10.1007/ s11101-009-9147-3. [11] Carvajal-Millan E, Landillon V, Morel M-H, Rouau X, Doublier J-L, Micard V. Arabi‐ noxylan gels:  impact of the feruloylation degree on their structure and properties. Biomacromolecules. 2004;6(1):309-317. DOI: 10.1021/bm049629a. [12] Hopkins MJ, Englyst HN, Macfarlane S, Furrie E, Macfarlane GT, McBain AJ. Degra‐ dation of cross-linked and non-cross-linked arabinoxylans by the intestinal microbio‐ ta in children. Applied and Environmental Microbiology. 2003;69(11):6354-6360. DOI: 10.1128/aem.69.11.6354-6360.2003. [13] Grootaert C, Delcour JA, Courtin CM, Broekaert WF, Verstraete W, Van de Wiele T. Microbial metabolism and prebiotic potency of arabinoxylan oligosaccharides in the

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human intestine. Trends in Food Science & Technology. 2007;18(2):64-71. DOI: 10.1016/j.tifs.2006.08.004. [14] Carvajal-Millan E, Berlanga-Reyes C, Rascón-Chu A, Martínez-López AL, MárquezEscalante JA, Campa-Mada AC, Martínez-Robinson KG. In vitro evaluation of arabi‐ noxylan gels as an oral delivery system for insulin. Materials Research Society. 2012. DOI: 10.1557/opl.2012. [15] Carvajal-Millán E, Rascón-Chu A, Márquez-Escalante JA. Método para la obtención de goma de maíz a partir del líquido residual de la nixtamalización del grano de maíz. Mexican Patent 2005;No. 278,768. [16] Morales-Ortega A, Carvajal-Millan E, López-Franco Y, Rascón-Chu A, Lizardi-Men‐ doza J, Torres-Chavez P, Campa-Mada A. Characterization of water extractable ara‐ binoxylans from a spring wheat flour: rheological properties and microstructure. Molecules. 2013;18(7):8417-8428. DOI: 10.3390/molecules18078417. [17] Martínez-López AL, Carvajal-Millan E, Rascón-Chu A, Márquez-Escalante J, Martí‐ nez-Robinson K. Gels of ferulated arabinoxylans extracted from nixtamalized and non-nixtamalized maize bran: rheological and structural characteristics. CyTA - Jour‐ nal of Food. 2013;11(sup1):22-28. DOI: 10.1080/19476337.2013.781679. [18] Vansteenkiste E, Babot C, Rouau X, Micard V. Oxidative gelation of feruloylated ara‐ binoxylan as affected by protein. Influence on protein enzymatic hydrolysis. Food Hydrocolloids. 2004;18(4):557-564. DOI: 10.1016/j.foodhyd.2003.09.004. [19] Morales-Ortega A, Carvajal-Millan E, Brown-Bojorquez F, Rascón-Chu A, TorresChavez P, López-Franco Y, Lizardi-Mendoza J, Martínez-López AL, Campa-Mada AC. Entrapment of probiotics in water extractable arabinoxylan gels: rheological and microstructural characterization. Molecules. 2014;19(3):3628-3637. DOI: 10.3390/mole‐ cules19033628. [20] Carvajal-Millan E, Rascón-Chu A, Márquez-Escalante JA, Micard V, Ponce de León N, Gardea A. Maize bran gum: Extraction, characterization and functional proper‐ ties. Carbohydrate Polymers. 2007;69(2):280-285. DOI: 10.1016/j.carbpol.2006.10.006. [21] Niño-Medina G, Carvajal-Millan E, Lizardi J, Rascón-Chu A, Gardea A. Feruloylated arabinoxylans recovered from low-value maize by-products In: Jones CE, editor. En‐ cyclopedia of Polymer Research: Nova Science Publishers, Inc.; 2011. p. 1401-1416. [22] Iravani S, Fitchett CS, Georget DMR. Physical characterization of arabinoxylan pow‐ der and its hydrogel containing a methyl xanthine. Carbohydrate Polymers. 2011;85(1):201-207. DOI: 10.1016/j.carbpol.2011.02.017. [23] Sárossy Z, Tenkanen M, Pitkänen L, Bjerre A-B, Plackett D. Extraction and chemical characterization of rye arabinoxylan and the effect of β-glucan on the mechanical and barrier properties of cast arabinoxylan films. Food Hydrocolloids. 2013;30(1):206-216. DOI: 10.1016/j.foodhyd.2012.05.022.

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[24] Kačuráková M, Wellner N, Ebringerová A, Hromádková Z, Wilson RH, Belton PS. Characterisation of xylan-type polysaccharides and associated cell wall components by FT-IR and FT-Raman spectroscopies. Food Hydrocolloids. 1999;13:35-41. [25] Egüés I, Stepan AM, Eceiza A, Toriz G, Gatenholm P, Labidi J. Corncob arabinoxylan for new materials. Carbohydrate Polymers. 2014;102:12-20. DOI: 10.1016/j.carbpol. 2013.11.011. [26] Mussatto SI, Dragone G, Roberto IC. Ferulic and p-coumaric acids extraction by alka‐ line hydrolysis of brewer's spent grain. Industrial Crops and Products. 2007;25(2): 231-237. DOI: 10.1016/j.indcrop.2006.11.001. [27] Berlanga-Reyes CM. Carvajal-Millán E, Caire Juvera G, Rascón-Chu A, Marquez-Es‐ calante JA, Martínez-López AL. Laccase induced maize bran arabinoxylan gels: struc‐ tural and rheological properties. Food Science and Biotechnology. 2009;18:1027-1029. [28] Martínez-López AL, Carvajal-Millan E, Lizardi-Mendoza J, López-Franco YL, Ras‐ cón-Chu A, Salas-Muñoz E, Barron C, Micard V. The peroxidase/H2O2 system as a free radical-generating agent for gelling maize bran arabinoxylans: rheological and structural properties. Molecules. 2011;16(12):8410-8418. DOI: 10.3390/mole‐ cules16108410. [29] Ross-Murphy SB. Rheological characterisation of gels. Journal of Texture Studies. 1994;26:391-400. [30] Ayala-Soto FE, Serna-Saldívar SO, Pérez-Carrillo E, García-Lara S. Relationship between hydroxycinnamic profile with gelation capacity and rheological properties of arabinoxylans extracted from different maize fiber sources. Food Hydrocolloids. 2014;39:280-285. DOI: 10.1016/j.foodhyd.2014.01.017. [31] Berlanga-Reyes C, Carvajal-Millan E, Niño-Medina G, Rascón-Chu A, RamírezWong B, Magaña-Barajas E. Low-value maize and wheat by-products as a source of ferulated arabinoxylans. In: Einschlag, editor. Croatia: InTech; 2011. [32] Paes G, Chabbert B. Characterization of arabinoxylan/cellulose nanocrystals gels to investigate fluorescent probes mobility in bio inspired models of plant secondary cell wall. Biomacromolecules. 2012;13:206-214. [33] Marquez-Escalante J, Carvajal-Millan E, López-Franco Y, Lizardi-Mendoza J, Valen‐ zuela-Soto E, Rascón-Chu A, Faulds. Gels of water extractable arabinoxylans from a bread wheat variety: swelling and microstructure. Plants and Crop - The Biology and Biotechnology Research. iConcept Press. Retrieved from http://www.iconcept‐ press.com/books/plants-and-crop--the-biology-and-biotechnology-research/ BK056-0101/.

Chapter 5

Perspectives on Biological Treatment of Sanitary Landfill Leachate Andreja Žgajnar Gotvajn and Aleksander Pavko Additional information is available at the end of the chapter http://dx.doi.org/10.5772/60924

Abstract Landfilling, one of the prevailing worldwide waste management strategies, is presented together with its benefits and environmental risks. Aside from biogas, another non-avoidable product of landfilling is landfill leachate, which usually contains a variety of potentially hazardous inorganic and organic compounds. It can be treated by different physico-chemical and biological methods and their combina‐ tions. The composition and characteristics of landfill leachate are presented from the aspect of biotreatability. The treatment with activated sludge, mainly consisting of bacterial cultures under aerobic and anaerobic conditions in various reactor systems, is explained, including an extensive literature review. The potential of fungi and their extracellular enzymes for treatment of municipal landfill leachates is also presented, with a detailed review of the landfill leachate treatment studies. The future perspec‐ tives of biological treatment are also discussed. Keywords: Activated sludge treatment, biotreatability, fungal treatment, landfill leachate

1. Introduction Landfilling is still widely accepted and used in any waste management strategy, but it can constitute a hazard for the environment. This method generally offers lower cost of operation and maintenance when compared to other methods, such as incineration. Besides households and urban activities, the industry is directly associated with the production of large amounts

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of solid wastes. Several methodologies and strategies have been developed for the integrated management of these wastes. They start with pollution prevention, waste minimization (zero waste), reuse of products or their parts, as well as material and/or energy recovery. But in spite of all environmental policies, the majority of municipal and industrial wastes still end up at the landfill and the amount of deposited wastes is significant worldwide. Landfill still accounted for nearly 40% of municipal waste treated in the European Union in 2010. In the 25 countries of the European Union, 502 kg of municipal waste was generated per person in 2010, while 486 kg of municipal waste was treated per person: 38% was landfilled, 22% incinerated, 25% recycled, and 15% composted. In the deposited wastes, organics are still present even after thorough waste separation, mainly due to the dirty packages and other remains that could not be completely separated; thus, microbial processes dominate the stabilization of the waste and lead to the generation of the landfill gas, and dictate the amount and composition of the leachate. Landfill leachate is defined as wastewater formed due to precipitation, deposited waste moisture, and water, formed within the body of the landfill. Untreated leachates can permeate groundwater or mix with surface waters and contribute to the pollution of soil, ground water, and surface waters. Careful site management can reduce the quantity and increase the purity of the formed leachate, but it cannot completely eliminate it. Its composition is therefore site- and timespecific, based on the characteristics of deposited solid wastes, physico-chemical conditions, rainfall regime that regulates moisture level, and landfill age. Even within a single landfill site, variability is frequently evident [1, 2, 3]. Significant components of leachate at the beginning of landfill operation are heavy metals and degradable organics, while persistent organic pollutants usually appear later as a result of biotic (i.e., living components that constitute an ecosystem) and abiotic (i.e., non-living chemical and physical components that affect living organisms and the performance of ecosystems) processes in the system. Among these sub‐ stances are several compounds classified as potentially hazardous: bio-accumulative, toxic, genotoxic (chemical compounds that damage the genetic information within a cell causing mutation that may lead to cancer), and they could have endocrine disruptive effect [2]. Hazardous substances from the leachate should be caught and removed properly, to avoid spreading in the receiving environment. Efficient treatment methods must be matched to the actual characteristics of a particular leachate and they could vary with time. Often, biological processes are employed if biotreatability in terms of low toxicity and at least moderate biodegradability of the leachate is indicated [2, 4]. Biodegradability of the wastewaters and also leachates is usually determined using various non-standardized laboratory or pilot-scale long-term tests with activated sludge as the source of active microorganisms [5]. Toxicity tests must be accomplished prior to the biodegradability determination to assess the impact of landfill leachate components on microorganisms of the aerobic or anaerobic activated sludge. Biodegradability assessment of leachates usually starts with the determination of ready biodegradation in common environmental conditions, it is upgraded with the assessment of biodegradation potential in an inherent biodegradability assessment test under optimal conditions, and it is finally concluded with a simulation of biodegradation in the wastewater treatment plant. All of the mentioned tests are based on the

Perspectives on Biological Treatment of Sanitary Landfill Leachate http://dx.doi.org/10.5772/60924

measurement of summary parameters, such as chemical oxygen demand (COD) or dissolved organic carbon (DOC) removal, O2 consumption, etc. Inherent biodegradability assessment tests provide data on adsorption potential of the sample to the activated sludge and allow estimation of its impact on the biological wastewater treatment plant. Preliminary estimations should then be verified in a laboratory or a pilot-scale aerobic treatment plant to determine the actual impact of the wastewater on the activated sludge processes [5-7]. The connection between the biodegradation and changes in toxicity of the sample represents the stabilization study, where leachate is diluted in a batch reactor to avoid significant toxicity, and it is aerated and stirred until the biodegradation reaches the final plateau. Among other parameters, toxicity is monitored during biodegradation by using the appropriate method. Stabilization (ageing) allows us to assess the toxic fraction as permanent or biodegradable [8-10]. After the complete examination of the landfill leachate characteristics, the appropriate treatment process should be considered. In the case of significant biodegradability, various biological processes could be involved. Biological treatment is reliable, simple, highly costeffective, and provides many advantages in terms of biodegradable and nitrogen and phos‐ phorous compounds removal [11, 12]. It can be accomplished with microorganisms in different types of reactors, in aerobic and anaerobic conditions. Classical systems with activated sludge, sequencing batch reactor (SBR), biofilters, membrane bioreactors, as well as up-flow anaerobic sludge blanket processes and fluidized bed reactors are often considered [13]. In the last few decades, researchers also confirmed the great potential of white rot fungi for removal of hazardous as well as toxic pollutants. They produce various extracellular ligninolytic enzymes, including laccase (Lac) and manganese peroxidase (MnP), which are involved in the degra‐ dation of lignin in their natural lignocellulosic substrates [14], and according to the literature data [15], offer also an interesting potential for the landfill leachate treatment. A considerable amount of work has been done in the field of landfill leachate biological treatment in the past decades. But the strict implementation of environmental legislative demands and the ageing of existing landfills put pressure on managers and operators of landfills to implement more efficient processes for landfill leachate treatment. Nevertheless, the results, obtained during decades of research, indicate some future research guidelines.

2. Solid waste management The industrial and economic growth of many countries around the world has resulted in a rapid increase in industrial and municipal solid waste generation [1]. Many countries are still trying to implement separate collection and disposal of waste based on the Reduce-RecycleReuse principle, which should lead to zero waste management practice in the future. However, the improved waste management strategy, based on the waste management hierarchy pyramid (Figure 1) clearly pointed out the disposal of solid wastes, such as landfilling, as the least favorable option leading to severe environmental impact and degradation. It also results in a loss of natural resources [16]. Different recovery options, such as energy recovery, result in the utilization of energy potential of the waste, while recycling is even preferred due to the

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recovery of materials. In both cases, some loss of material and/or energy is noticed, and thus the differences regarding the impacts on the environment are relatively small [17]. The measures of waste prevention eliminate the need for the above-mentioned measures, thus lowering the impact that the waste has on the environment. The waste prevention measures can be any form of reducing the quantities of materials used in a process, or any form of reducing the quantity of harmful materials that may be contained in a product. At the same time, the replacement of hazardous materials should also be considered. Prevention and minimization also include processes or activities that avoid, reduce, or eliminate the waste at its source, or result in its reuse or recycling [18]. In recent years, these goals were incorporated in European environmental policy to make a resource-efficient Europe, while cutting off the burden of wastes and emissions. Effective implementation of these waste policies demands an understanding of what has been achieved so far and a set-up of targets and broader goals for the future. The first EU legislation covering the generation and treatment of waste was introduced in 1975, and now, there are more than 20 legislative documents currently in force on waste management [19-22]. Municipal solid waste is defined differently in different European states, but according to Eurostat, “Municipal waste is mainly produced by households, though similar wastes from sources such as commerce, offices, and public institutions are included. The amount of municipal waste generated consists of waste collected by or on behalf of municipal authorities and disposed of through the waste management system” [23]. Other wastes, similar in nature and composition, but collected by the private sector, are also included in this definition. In the EU Landfill Directive, municipal solid waste is defined broader as the “Waste from house‐ holds, as well as other waste which, because of its nature or composition, is similar to waste from households” [24]. This seems to be the definitively more appropriate definition, referring to the type of the waste and not to who produced or collected it [25]. However, due to different definitions and methodologies used across EU countries, the efficiency of improved waste prevention in the last decade (2001-2010) is not easy to estimate. The average EU-27 value still varies around 500 kg for wastes generated per capita. 21 countries generated more municipal wastes per capita in 2010 than in 2001, and 11 cut per capita municipal waste generation [25]. On the other hand, there was a clear shift up the waste hierarchy, from landfilling to energy recovery and recycling. In this period of time, landfilling decreased by almost 41 million tons, whereas incineration increased by 15 million tons, and recycling grew by 28 million tons within EU-27. In the study where, together with the EU-27 countries, Croatia, Iceland, Norway, Switzerland, and Turkey were also taken into account, the decrease of landfilling in 2001-2010 is evident in Table 1 [25]. Each country can be included in several waste management categories, so the total number of countries is sometimes more than 32. When Table 1 is considered, the shift in recycling is noticeable; however, in 2010, in 50% of the countries landfill still presented more than 50% of municipal waste. The data on the trends in the recycling of materials and biowaste could be used to evaluate changes in the composition of deposited solid wastes. This parameter is also affected by the implementation of waste hierarchy, not only by the quantity of the deposited wastes. The total recycling rate increased between 2001 and 2010, mainly due to the fact that many countries have increased

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119

recycling of materials such as glass, paper and cardboard, metals, plastic, and textiles. Eight out of 32 investigated countries increased their material recycling rate by more than 10%, and 11 countries achieved an increase of 5%-10% [25].

Most desirable option

Prevention

Reuse 

Recycling 

 Energy recovery 

Least desirable option 

Disposal 

Figure 1: Waste management hierarchy [18].

Figure 1. Waste management hierarchy [18].

together with the EU-27 countries, Croatia, Iceland, Norway, Switzerland, and Turkey were also taken into account, the decrease of landfilling in 2001–2010 is evident in Table 1 [25]. Each country can be included in several waste management categories, so the total number of countries is sometimes more than 32. When Table 1 is considered, the Waste management Number of countries shift in recycling is noticeable; however, in 2010, in 50% of the countries landfill still presented more than 50% of municipal waste. The data on the trends in the recycling of materials to evaluate changes in Year 2001 and biowaste could be used Year 2010 the composition of deposited solid wastes. This parameter is also affected by the implementation of waste hierarchy, not > 25 % of Recycling 11 rate increased between 2001 and 16 2010, mainly due to only by the quantity the deposited wastes. The total recycling the fact that many countries have increased recycling of materials such as glass, paper and cardboard, metals, plastic, > 25 % Incineration 8 10 and textiles. Eight out of 32 investigated countries increased their material recycling rate by more than 10%, and 11 > 50 % 22 19 countries achieved anLandfilling increase of 5%–10% [25]. > 75 % Landfilling

Waste management

17

11

Number of countries Year 2001 Year 2010 Table 1. Number of countries at different levels of municipal waste management in 2001 and 2010 [25]. > 25 % Recycling 11 16 > 25 % Incineration 8 10 In contrast to material recycling, biowaste22 recycling is not comparably efficient over the same > 50 % Landfilling 19 > 75 % of Landfilling 11 period time (EEA, 2013); one country17increased its municipal waste-derived biowaste

recycling by more than 10%, and only six of them improved it by 5%-10%. Eighteen countries even sustained a very low level of biowaste recycling (0%-10% of municipal waste generated). This could be a consequence of several affected by the methodology of data collection In contrast to material recycling, biowaste recycling is notfactors comparably efficient over the same period of time (EEA, 2013); one country increased municipal The waste-derived by more than 10%, and only six of them improved and theits legislation. fact is thatbiowaste materialrecycling and biowaste recycling potential depends on the it by 5%–10%. share Eighteen countries sustained very lowmunicipal level of biowaste recycling (0%–10%the of biowaste municipal waste of each waste even type in the totalacollected waste. In most countries, generated). This could be a consequence of several factors affected by the methodology of data collection and the lower than recycling the material recycling potential represents legislation. Therecycling fact is thatpotential material is and biowaste potential depends on thebecause share ofbiowaste each waste type in the total a smaller proportion the total municipal waste [25]. This leads toisthe conclusion thatmaterial different collected municipal waste. In most of countries, the biowaste recycling potential lower than the recycling Table 1. Number of countries at different levels of municipal waste management in 2001 and 2010 [25].

potential because biowaste represents a smaller proportion of the total municipal waste [25]. This leads to the conclusion that different policy instruments should be implemented and combined in different countries to achieve maximal impact in terms of municipal waste management.

3.

Landfilling of Municipal Waste

Currently, the deposit in a landfill is still the most widely used method for municipal solid waste disposal within almost all European states (Table 1). The landfill can be considered as a complex environment or even a biochemical reactor,

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policy instruments should be implemented and combined in different countries to achieve maximal impact in terms of municipal waste management.

3. Landfilling of municipal waste Currently, the deposit in a landfill is still the most widely used method for municipal solid waste disposal within almost all European states (Table 1). The landfill can be considered as a complex environment or even a biochemical reactor, where many interacting physical, chemical, and biological processes take place. The degradation process of municipal solid wastes in landfills is a long-term event. A major problem regarding disposal of wastes is the lack of available landfilling sites, as well as the production of landfill leachates and biogas, consisted mainly of carbon dioxide and methane, which has 28 times higher global warming potential than CO2 in a 100-year cycle [26]. In Slovenia, for example, net emissions of green‐ house gases due to municipal waste management have been decreasing constantly since 1999 [27]. This could also be the consequence of a better municipal waste management (e.g., recycling of biowaste), resulting in lower biodegradable fraction landfilled. It can be assumed that the direct emissions of GHCs from landfilling will continue to decrease in the coming years, but for several years ahead, considerable amounts of greenhouse gases will continue to be emitted from landfills because biowaste landfilled in the previous years will continue to generate methane for several decades [27]. However, with appropriate entrapment and utilization, biogas is usually efficiently exploited for energy purposes at the site, while leachates could pose a serious risk for nearby soil, surface, and underground waters [28]. At the sites, where there is no need for energy or where the methane content is very low, methane is flared to avoid its migration in the atmosphere. Landfill top covers or so-called biocovers are often used at landfills to reduce methane emissions. They optimize environmental conditions for development of methanotrophic bacteria to help oxidize any fugitive methane. Biocovers are usually spread over the entire surface of the landfill and they are made of compost, dewatered sewage sludge, or other waste material. Landfills with gas collection and recovery systems had a methane recovery efficiency of 41%-81%. Methane emissions could range from 2.6 kg h-1 to 60.8 kg h-1, with the lowest emissions from the small and old landfills and the highest emissions from the larger landfills [29]. 3.1. Landfill leachate Appropriate management can reduce the quantity and quality of the leachate, but it cannot be completely eliminated. Landfill leachate is generated as a mixture of rainwater percolation through the wastes, water produced from the (bio) degradation of wastes, and the water present in the wastes at the time of deposition [2]. Its composition is based on the type and the amount of waste deposited, as well as its maturity, and the construction of the landfill site [4]. Main sequential and distinct consecutive stages involved in landfill stabilization are [30, 31]: 1.

Aerobic phase, characterized by processes enabled by oxygen present. In this acclimation phase, sufficient moisture develops to support active microbial communities. Initial

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changes in compounds occur in order to establish the appropriate conditions for further biochemical degradation. 2.

Hydrolysis and fermentation stage, where complex molecules are broken into smaller fragments and an aerobic environment is transferred to an anaerobic one, the amount of entrapped oxygen is drastically reduced and the reducing conditions occur. The main electron acceptors are nitrates and sulfates.

3.

Anaerobic acetogenic stage, characterized by continuous hydrolysis of solid wastes, preceded by the production of volatile fatty acids at high concentrations. Low hydrogen levels promote the activity of methanogenic bacteria, which produce methane and carbon dioxide from organic acids. This stage could be recognized by a high concentra‐ tion of metals in the leachate, due to their increased mobility because of the lower pH, even below 4.

4.

Anaerobic methanogenic stage, where significant methane production is evident. pH is again increased to 7-8, due to the degradation of intermediate acids and the buffering capacity of bicarbonates. The concentration of heavy metals is reduced again, due to their complexation and precipitation. During this stage, the mesophilic bacteria, which are active in temperatures of 30°C-35°C, and thermophilic bacteria, active in the range of 45°C-65°C, dominate the microbial population.

5.

Maturation stage, as the final stage of landfill stabilization, could be recognized by low microbial activity due to degradation of biodegradable fraction and limiting impact of nutrients. As a result, the methane production decreases, as does the amount of pollutants in the leachate, which usually stays at a constant level. Slow degradation activity of the resistant organic pollutants can be observed by the production of humic and fulvic substances.

The duration of each phase is dependent upon many factors, and the development and activity of microorganisms are dependent upon sustaining appropriate conditions. Landfill leachates are characterized by high concentrations of numerous toxic and carcinogenic chemicals, heavy metals, and organic as well as inorganic matter. Among the organic compounds detected in the landfill leachate, the main compounds are different hydrocarbons, esters, alcohols, and ketones, as well as aromatic and heterocyclic compounds [32]. Additionally, the leachates can also be contaminated with bacteria, including aerobic, psychrophilic and mesophilic bacteria, coliform and fecal coliforms, spore-forming-bacteria, and with numerous fungi [32, 33]. Typical concentrations of landfill leachate compounds as a function of landfill age and stabilization are presented in Table 2 [30, 31, 34]. Concentrations of organic compounds, expressed as COD and biological oxygen demand 5day test (BOD5), in the young leachate are high (aerobic and acid formation phase), while leachates from stabilized landfills (methane formation and maturation phase) contain lower levels of organic matter [28]. Several authors have reported that in young landfills, COD could vary from 5,000 to even more then 60,000 mg L-1, with BOD5 starting from 3,000 L-1 to even 40,000 mg L-1 [35]. Those values are significantly lower in leachates from mature landfills, as a result of biotic stabilization processes in the body of the landfill [36]. The total quantity of

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the produced landfill leachate can be estimated by using empirical data based on flow measurements, or by using water mass balance between precipitation, evapotranspiration, surface runoff, and capacity of moisture storage. Waterproof covers and different covering liners contribute a lot to the reduction of landfill leachate quality, but they cannot completely reduce it. Both parameters (leachate quality and quantity) affect the attempts of uniform design of leachate treatment systems. The optimal treatment solution may change over time because of the changeable quality and quantity of the leachate, and the development of new technol‐ ogies and the legislation [37]. Parameter

Aerobic phase

pH (/)

Acid formation

Methane formation

phase

phase

Maturation phase

6.7

4.7–7.7

6.3–8.8

7.1–8.8

COD (mg L )

480–18,000

1500–71,000

500–10,000

0.5

0.5–1.0

0.5–1.0

100

99%*

DOC

90%

Laboratory

⋅ Magnetic ion exchange

⋅ Aerated lagoon

Pilot

⋅ Solar photo-Fenton

Young/

After lagooning TN

⋅ Conventional biological

[62]

[63]

Granular Reactor

⋅ FeCl3 coagulation

Reference

efficiency

[64]

[61]

56%–90%*

Biodegradability

Increased

Mature/

COD

97.4%

Stabilized

TOC

92.3%

⋅ SBR, mixed with sewage: anoxic-

BOD5

94.4%

aerobic-anoxic conditions

SS

97.5%

⋅ Sand and carbon filtration

NH3-N

99.2%

Mature/

COD

87%

Stabilized

BOD5

90%

NH3-N

43%–79%*

WWTP (with nitrification/ denitrification) ⋅ Agitation/stripping

Laboratory

⋅ FeSO4 coagulation

⋅ TiO2/UV photolysis

Laboratory

⋅ Bioreactors with various inoculums (raw leachate/soil

[66]

[65]

extract/activated sludge) ⋅ Coagulation/flocculation

Laboratory

Mature/

Toxicity:

⋅ Fenton

Pilot

Stabilized

-respirometry

Remains low

COD

89%

Biodegradability

Increased

⋅ Solar photo-Fenton

[67]

*...Depends upon the combination of treatment processes and landfill leachate sample characteristics. TP = Total phosphorous TN = Total nitrogen SS = Suspended solids COD = Chemical Oxygen Demand DOC = Dissolved Organic Carbon OM = Organic Matter DOM = Dissolved Organic Matter WWTP = wastewater treatment plant Table 4. Some of the recently investigated combinations for treatment of heavily polluted landfill leachates [61-67].

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The TIE methodology uses the responses of the test organisms to detect the presence of toxic substances in the sample before and after the samples are subjected to a series of physical and chemical treatments. This combination of physical/chemical manipulations of toxic samples, followed by the toxicity testing, allows one to isolate and identify the problematic group of compounds [50, 54, 55]. The most often used procedures are listed in Table 3. Results could be efficiently utilized to set up the appropriate treatment procedure for the particular wastewater, because it can be clearly estimated which treatment method is efficient in the removal of the particular group of pollutants and where the toxicity of the wastewater comes from [56]. Usually these simple, cost-effective methods could reduce the need for a complex and detailed characterization of the wastewater before setting up treatment procedure. However, they have to be designed and performed with caution (blank sample) to avoid any impact of the applied method (pH manipulation, addition of chemicals, etc.) to the final characteristics of the sample. According to the results of the methods described in Table 3, suitable treatment methodology could be set up. If, for example, the wastewater contains a significant fraction of a nonbiodegradable organic fraction (determined by biodegradability testing) and it contains a lot of oxidizable organics (proved by oxidation experiment), one of the advanced oxidation processes would seem to be to be the most viable treatment option. On the other hand, if it contains organics that are able to mineralize almost completely in the biodegradability test, and it contains a lot of ammonia, one of the biological treatments involving nitrification/ denitrification would seem to be to be the best choice. It can be clearly concluded that, in the case of municipal landfill leachates, a technically and economically viable methodology for the effective treatment has yet to be designed. The available options are similar to those used in the treatment of industrial wastewaters, involving a combination of physical, chemical, and biological processes. Primarily due to their low costs, the biological processes, in their various forms according to redox regime (aerobic, anaerobic, anoxic), a type of biomass (a mixed or a pure bacterial culture, fungi, etc.) and a biomass fixation (dispersed, attached), remain the most widely implemented type of treatment processes [2, 58, 59]. However, a combination of biological and physico-chemical processes is usually employed for heavily polluted leachates. Many examples of efficient treatment combinations could be found in literature. As presented in [60], an aerobic biological treatment, a chemical coagulation, an advanced oxidation process (AOP), and some combined treatment strategies were compared. Laboratory experiments were done with 200 mL samples in a glass vessel. The efficiency of these treatment procedures was evaluated by analyzing the COD and color removals. In the extended aeration process, the maximum COD and color removals were 36% and 20%, respectively. They could be achieved during the optimum retention time of 7 days. Chemical coagulation with an optimum aluminum sulphate dose of 15,000 mg/L at pH = 7.0, gave the maximum COD and color removals of 34% and 66%, respectively. Using Fenton oxidation process at optimum pH = 5.0 and optimum dosages of reagents, with H2O2/Fe2+ molar ratio of 1:3, the highest removals of COD and color were 68% and 87%, respectively. The combined treatment, the extended aeration followed by Fenton oxidation, was found to be the most suitable.

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Some additional, recently investigated and proposed treatment designs are presented in Table 4. A large-scale multistage treatment system was also designed for the treatment of a mature raw landfill leachate [61]. The system consisted of an activated sludge biological oxidation (ASBO) reactor for aerobic and anoxic conditions (volume 3.3 m3) and a solar compound parabolic collector (CPC) for photo-Fenton process (total collector surface 39.52 m2 and illuminated volume 482 L). The raw leachate was characterized by a high concentration of humic substan‐ ces, representing 39% of the DOC content and high nitrogen content, mostly in the form of ammonium nitrogen. In the first biological oxidation step, a 95% removal of total nitrogen and a 39% mineralization in terms of DOC were achieved. The following photo-Fenton reaction led to the depletion of humic substances > 80% of low-molecular-weight carboxylate anions > 70% and other organic micropollutants, thus resulting in a total biodegradability increase of > 70%. The neutralized photo-bio-treated leachate was finally treated with the second stage biological oxidation, where the rest of biodegradable organic carbon and nitrogen content were eliminated. This way, a high efficiency of the overall treatment process was achieved.

4. Biological treatment of the landfill leachate 4.1. Treatment with activated sludge Biological treatment has become one of the most often used treatment processes; it is the most common method for the removal of organic, nitrogen, or phosphorus components from wastewaters. One of the main reasons for the selection of this process is its capability to achieve high elimination efficiency of these pollutants, and at the same time, it is relatively less expensive than physico-chemical or chemical processes. The pollution is completely destroyed to the level of non-hazardous, simple products, and not only transformed into another form. Nowadays, it is used not only for the treatment of sewage, but also for the removal of different xenobiotics such as pharmaceuticals, personal care products, and cleaning agents from the sewage and the heavily polluted industrial wastewaters and landfill leachates [12]. Biological degradation of pollutants is caused by the metabolic activity of microorganisms, in particular by the bacteria and fungi that live in natural environments. However, its efficiency is strongly reduced in the presence of refractory or inhibitory compounds in wastewaters, typical also for mature landfill leachates [33, 66, 68]. To achieve good removal efficiency, high BOD/COD ratio is recommended (>0.5) [13]. When biological treatment is discussed, mainly microorganisms that grow in a controlled environment through a complex sequence of biochemical reactions, forming the vital steps of their metabolic activities are considered [13]. The prevailing species are the saprotrophic bacteria, there is also an important protozoan flora present, composed mainly of amoebae, Spirotrichs, Peritrichs including Vorticellids, and a range of other filter-feeding species. Fungi could also contribute to the diversity of present populations. Other important constituents include motile and sedentary rotifers. The most important seems to be the bacteria, found in all types of treatment processes. The nature of the population changes continually, in response

Perspectives on Biological Treatment of Sanitary Landfill Leachate http://dx.doi.org/10.5772/60924

to variations in the composition of the wastewaters and to environmental conditions [69]. Generally, biological treatment of wastewaters involves bacterial community in aerobic and anaerobic conditions, which could be dispersed or attached, in small flocks, granulated or forming biofilms, while treatment with fungi and their enzymes also lately received more attention [70]. However, fungal treatments have not yet found a wider recognition, due to the difficulty in selecting organisms that are able to grow and remain active in the actual waste‐ water [71]. 4.1.1. Treatment with activated sludge under aerobic conditions The most often applied processes for biological treatment are aerobic. In an aerobic environ‐ ment (concentration of dissolved oxygen >2 mg L-1), organic matter is used as a food source for microorganisms. The suspended organics are removed by entrapment in the biological activated sludge flocks. The colloidal organics and a small amount of soluble organics are also partially adsorbed and entrapped by the sludge flocks. Therefore, approximately up to 85% removal of the total COD can be achieved after 10 min to 15 min of retention time. The remaining degradable soluble organic fraction undergoes biological reactions [11, 13]. A portion of organic compounds (about 50% of organic carbon) is oxidized to CO2 and H2O, and the rest of it is incorporated into a new biomass. Approximately about 60% of the energy content in wastewater organics is consumed for synthesis of the new biomass and the rest represents reaction heat loss [13]. At the same time, an efficient removal of ammonium nitrogen should also be achieved to protect the sensitive water bodies from eutrophication [12, 54]. A combination of aerobic and anoxic environment is necessary for the accomplishment of organic and nitrogen pollutants’ removal from wastewater. On the other hand, the combina‐ tion of anaerobic and aerobic environment is required to biologically remove phosphates from wastewater, so the systems could not always be characterized as completely aerobic or anaerobic. The upgrading of biological processes, from removal of carbonaceous organics to the nitrogen and phosphorus removal, significantly impacted the system configuration. Not only do the system configuration and its operation increased in complexity, but also the new legislative demands on effluent quality have to be met. Thus, the system must be welldesigned, optimized, and operated at its optimum in order to fulfill these criteria [72]. Some of the bioreactors, also applicable for aerobic treatment using microorganisms, are summar‐ ized in Table 5. Aerobic biological systems, based on suspended-growth biomass, have been widely studied and also applied [2]. Also recently, attached-biomass systems have been developed, such as moving bed bioreactors (MBBR) and with different options of biofilters. A promising alternative also are membrane bioreactors (MBR), which represent an advanced biological treatment process, replacing a secondary clarifier in activated sludge process for removal of biomass with membrane module. It can be incorporated as an internal or an external unit of aeration basin to achieve better effluent quality, process stability, increased biomass retention time, and low sludge production [73]. Some of the systems where biological treat‐ ment represents the most effective stage will be briefly overviewed and discussed in this chapter. It should be emphasized that sometimes it is difficult to distinguish between aerobic and anaerobic treatment plants, due to the fact that most of the systems apply a combination of different regimes (aerobic, anaerobic, anoxic) to achieve the optimal treatment efficiency.

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Systems with suspended-growth biomass

Systems with attached/immobilized biomass

Lagoons:

Biofilters:

⋅ Aerated

⋅ Tricking filters

⋅ Non-aerated

⋅ Submerged biological filters

Constructed wetlands:

Moving bed bioreactors (MBBR):

⋅ Horizontal system

⋅ Rotating biological contactors

⋅ Vertical system

⋅ Suspended carrier biofilm reactors

Activated sludge (AS) systems: ⋅ Continuous flow reactors ⋅ SBR Membrane bioreactors (MBR): ⋅ External membrane module ⋅ Submerged/Immersed membrane module Table 5. Typical treatment system with aerobic microorganisms applied in landfill leachate treatment [2, 13, 73].

When aerated and non-aerated lagoons are discussed, together with artificial or natural wetlands, there is usually a combination of aerobic-anaerobic systems. The upper part is usually aerated, while the bottom part is anaerobic [74]. Such combination is well-illustrated in [75]. Four connected on-site lagoons were used for the treatment of mature landfill leachate in the aging methanogenic state. The landfill leachate contained a relatively low COD value (mean value 1,740 mg L-1) and a relatively high ammonium nitrogen concentration (mean value 1,241 mg L-1). The pH of the raw leachate was in the range of 7.0-8.0 and the temperature was 16.7ºC, which is higher than the mean ambient temperature (13.5ºC). Volumes of the lagoons varied from 60-80 m3. The leachate was mixed and aerated by compressed air (4-6 h) through diffuser pipes at the base of lagoons. The facultative aerobic system was obtained where sequential aerobic and anaerobic stages were maintained. The total COD removal was 75% in 56 days. Ammonium nitrogen removed 99%, while an average of 9 mg L-1 remained. However, the authors calculated that the conditions at the site could still be toxic to fish and the treated leachate should be diluted before its release into the environment. Due to aerobic conditions, nitrification occurred and the concentration of the nitrate was considerably higher in the effluent of the lagoons than in the influent; however, 80% of nitrogen was removed. The bacterial community profile also differed from one lagoon to another. The study by [76] compared efficiency of horizontal- and vertical-constructed wetlands for landfill leachate, containing 2,930-14,650 mg L-1 of COD, 170-4,012 mg L-1 of ammonium nitrogen, and 44-153 mg L-1 of ortophosphate-P. The experiments were run in a continuous mode, in three subsurface wetland systems; two of them operated in a vertical flow mode and one in a horizontal flow mode. The system was planted by Typha latifolia. Basins were 100 cm in length, 50 cm wide and 40 cm deep. The systems were filled with different heights of gravel and sand when zeolite was added in the third lagoon to increase its adsorption and ion exchange capacity. The leachate was introduced intermittently (10 min h-1) to assure hydraulic retention time of 8-12.5 days. In the vertical systems, COD removal was 15%-42%, while it

Perspectives on Biological Treatment of Sanitary Landfill Leachate http://dx.doi.org/10.5772/60924

reached up to 61% in the horizontal one. Ortophosphate-P removal in the vertical systems was 30%-83%, while in the horizontal one, it varied between 26.3 and 61.0%, depending on the climate conditions (a month of determination). The removal of NH4-N was better in vertical systems (36.8-67.4) in comparison to the horizontal one (17.8-49.0). The authors also presented the removal of heavy metals from the leachate. The concentration of Cr and Zn increased in the effluent, suggesting that they were washed out of the system. The iron removal decreased with time and was below 50%. The removal of lead varied between 30%-90%. It was concluded that the vertical system with zeolite layer was beneficial, especially in terms of ammonia removal. In the study of [77], the activated sludge system with sequencing batch laboratory reactor mode was employed for the treatment of landfill leachate, containing 4,298-5,547 mg L-1 of COD, 913-1,017 mg L-1 of BOD5, 13,971-17,421 mg L-1 of TDS, and 72-374 mg L-1 of NH4-N. Biomass was in the form of granules (0.36-0.60 mm, in cylindrical shape), the working volume of the reactor was 3 L in the operational mode of 12-h cycle (60 min of feeding, 640 min of aeration, 5 min of settling, and 5 min of effluent discharge). The reactor operated for three months with 182 cycles. The average COD removal was 82.8%-84.4%, while the ammonium removal efficiency reached 62±8%. The authors confirmed the high impact of initial ammonium N concentration on the performance of nitrification, which was confirmed by the simultaneous experiment with pretreated landfill leachate, where NH4+-N was reduced. To obtain a high removal of organics, a pretreatment in terms of ammonium removal was proposed. Removal of nutrients, especially ammonium nitrogen, from landfill leachates using a biological system was also studied by [78]. A batch reactor (V=150 L) was used for the treatment of homogenous mixture of old and freshly produced leachate from the body of the Tunisian landfill. The first treatment step involved an anoxic process preceded by aerobic processes. In the anoxic phase, COD reduction reached 46%, TOC was reduced to 65%, while ammonium N removal was 45%. Afterwards, the treated leachate was led to three aerobic submerged biological reactors, where a 7-day retention time was employed. As a result, high overall treatment efficiencies were obtained in BOD5, COD, and the NH4+-N removal was 95%, 94%, and 92%, respectively. MBR also gained a lot of attention lately for the treatment of landfill leachate [73]. They are often referred to as an efficient and robust alternative to other systems, in spite of their higher operational costs. They are essentially composed of two parts; i) a bioreactor dealing with the removal of organics and ii) a membrane module for the separation of the treated leachate and biomass. In comparison to the conventional activated sludge systems, they allow for complete retention of the biomass in the system; this way, the settling characteristics of the sludge are less important. As a result, the system can be operated at much higher concentrations of biosolids, up to 20 mg L-1, with very clean effluent. Additional important advantages are also higher loading rates, smaller volumes, lower production of excess sludge, and easier devel‐ opment of microorganisms with lower growth rates. MBR systems are usually designed as ultrafiltration or nanofiltration in a hollow fiber, plate, or frame; they could be in a flat or tubular configuration with continuous stirred tank reactors, plug-flow systems or sequencing batch reactors.

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A submerged MBR for the treatment of heavily polluted landfill leachate, containing 18,685 mg L-1 of COD and 310 mg N per liter was used in reference [3]. Biologically treated leachate was additionally filtrated, using nanofiltration and reverse osmosis. The biological stage was very effective, removing 89% of COD and 85% of the total Kljeldahl nitrogen (TKN). In MBR, polyeter sulfone ultrafiltration membrane was installed as a submerged module. A system with the volume of 4 L operated as SBR. The landfill leachate was fed daily (300 mL) and the lack of phosphorus was overcome by the addition of KH2PO4. The system reached steady-state operational mode after 4 months with 9 g L-1 of biomass. COD removal stabilized at 89%. Some portion of inert COD was entrapped in the reactor. Due to the unstable pH conditions, effective nitrification was not achieved; it was probably also reduced due to the air stripping of NH4-N at pH=8.6 and the consequent lack of the ammonium. However, sometimes it is impossible to distinguish clearly between systems with suspended and attached biomass, since the combination of advantages of both treatment systems is sometimes the most optimal to assure the environmentally and legislatively acceptable performance. A combination of a cross-flow MBR and MBBR for the treatment of stabilized leachate was, for example, used in reference [79]. The treatment was focused on the removal of ammonium N, present in stabilized landfill leachate up to 3,000 mg L-1. A combination of pure oxygen MBR and the subsequent MBBR was used for the nitrification and the denitrifi‐ cation, respectively. The volume of the membrane bioreactor was 500 L and it contained ultrafiltration ceramic membrane, while the volume of the MBBR was 540 L. Ammonium was oxidized only to a nitrite to reduce oxygen consumption for 25%, in comparison with the much higher oxygen consumption in the case of its complete conversion to nitrate. The authors obtained a 90% conversion of the ammonia N to nitrite with the sludge retention time of over 45 days. The system also enables up to 40% savings in the COD demand for nitrification. It was possible to oxidize more than 95% of total N inflow, and effluent ammonia concentrations were always below 50 mg L-1. Fixed bed filters offer a higher resistance to the toxic compounds and lower temperatures [80]. Aerated filters have been used for efficient treatment of landfill leachate [81]. In the study of [82], two systems using attached biomass were compared. Two MBBR systems using small free-floating polyurethane elements and granular activated carbon (GAC) as biomass carriers were set up. The laboratory SBR system had a volume of 6 L, and the inflow leachate contained NH4-N, COD, and BOD5 in the average values of 1,800 mg L-1, 5,000 mg L-1, and 1,000 mg L-1, respectively. According to the low BOD5/COD ratio=0.2 and pH=7.5, the landfill leachate could be characterized as a stabilized one. The study was conducted in two separate treatment cycles, the first one with cube-shaped polyurethane (30 g/reactor), while in the second one, 90 g of GAC (1,100 m2 g-1) was added. Nitrification and denitrification processes also occurred, but the need for additional external dosage of carbon source was emphasized. However, these processes efficiently and almost completely removed nitrogen, accompanied by the sufficient removal of COD (up to 81%), BOD5 (up to 90%), and turbidity. A bioreactor cascade with a submerged biofilm was also successfully used for a young landfill leachate treatment. Three reactors, each with the working volume of 18 L were used. The biofilm support was made from PVC synthetic fiber (57 m2 m-3). The reactors were inoculated

Perspectives on Biological Treatment of Sanitary Landfill Leachate http://dx.doi.org/10.5772/60924

with sludge from wastewater treatment plant and the biofilm consisted of various microor‐ ganisms, while bacterial groups such as Bacillus, Actinomyces, Pseudomonas, and Burkholderia genera were assumed to be responsible for the simultaneous removal of organic carbon and nitrogen. Bioreactor operated at a hydraulic retention time of 12 h, under organic loading charges 0.6 to 16.3 kg TOC m3 day-1. TOC removal rate varied between 65% and 97% and the total reduction of COD reached 92% without initial pH adjustment. The removal of total Kjeldahl nitrogen for loading charges of 0.5 kg N m3 day-1 reached 75%. However, pH increased during the experiments and caused biofilm separation and a decrease of attached solids concentration, which consequently reduced the carbon and nitrogen removal. When pH was adjusted to 7.5, nitrogen removal improved to 85% at a loading charge of 1 kg N m3 day-1 [83]. A pilot-scale submerged aerobic biofilter (SAB) with a working volume of 178 L and packed with polyethylene corrugated Racshig rings was used in reference [68]. Compressed air was used for aeration. It was continuously operated, with a hydraulic retention time of 24 h, and inoculated with activated sludge biomass. The co-treatment of domestic wastewater with a large content of biodegradable organic matter, and old landfill leachate with high total ammonium nitrogen (TAN) and extremely low BOD/COD ratios, was evaluated. The leachate volumetric ratios 2 and 5 v/v.% were tested. The best results were obtained at a volumetric ratio of 2 v/v.%, where 98% of the BOD, 80% of the COD and DOC, and 90% of the total suspended solids (TSS) were removed. Here, the poorly biodegradable organic matter in leachate was removed by partial degradation. When leachate was added at a volumetric ratio of 5 v/v.%, biodegradation of the low biodegradable organic matter was less efficient and its concentration decreased primarily as a result of dilution. The total ammonium nitrogen was mostly removed (90%) by nitrification. Generally, the treatment efficiency of the landfill leachate treatment is evaluated according to operational parameters of the investigated treatment plant, and on the basis of analyses of landfill leachate prior and after treatment. Non-specific parameters such as COD, BOD5, and DOC are routinely determined, as well as the metallic content, concentrations of different ions and some organic pollutants. However, the identification of particular problematic contami‐ nants (pesticides, personal care products, pharmaceuticals, endocrine disrupters, etc.) is difficult and impractical because of analytical limitations (low concentrations, complex extraction methods before analytical procedure, reactivity of the components, etc.), the uncertainty surrounding their bioavailability and the complexity of leachates [67], including possible additive, antagonistic, and synergistic effects of contaminants. As a result, there is a lack of studies dealing with complete determination of reduction of the landfill leachate hazardous environmental impact, which is most suitably determined by the battery of biotests [84]. Only a few studies addressing this problem could be found [56, 67, 70, 85]. 4.1.2. Treatment with activated sludge under anaerobic conditions The anaerobic biodegradation plays an important role in different environmental compart‐ ments, such as eutrophic lakes, soils, or sediments, while anaerobic digestion technology for waste and wastewater treatment as well as soil remediation is growing worldwide because of its economic and environmental benefits [12, 46]. The most favorable property of the anaerobic

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process is the biogas production, contributing to the renewable energy generation. Aerobic systems are suitable for the treatment of low strength wastewaters (BOD5 < 1,000 mg L-1), while anaerobic systems are suitable for the treatment of highly polluted wastewaters (BOD5 > 4,000 mg L-1), which is usually not the case when stabilized leachates are discussed. At the same time, after the initial aerobic (acetogenic) phase, landfills actually become anaerobic digesters by themselves [2]. The leachate produced after this phase has already been subjected to anaerobic digestion, so there is little additional treatment efficiency obtained by anaerobic treatment of such leachates. At the same time, anaerobic systems produce an effluent still containing a very high concentration of ammonium nitrogen, which needs a further aerobic stage, necessary to nitrify the anaerobic effluent to be suitable for a watercourse discharge [74, 86]. On the other hand, no such second stage is needed for the aerobic process [11]. Thus, the use of anaerobic-aerobic processes can lead to a reduction in operating costs compared to aerobic treatment alone, while simultaneously resulting in higher organic matter removal efficiency, efficient removal of nitrogen, and a lower waste sludge production. Anaerobicaerobic systems have received a great deal of attention over the past few decades due to their numerous advantages, not only with regard to the municipal wastewater, but also the sanitary landfill leachates (see Chapter 3.2). Aerobic-anaerobic systems incorporate advantages of both approaches [59]. They could be integrated bioreactors with or without physical separation of aerobic-anaerobic zones, the zones could be switched due to the sequencing mode of operation or they could employ combined culture of anaerobic and aerobic microorganism [74]. They usually achieve more than 70% of COD removal in a short hydraulic retention time (hoursdays). Anaerobic degradation of wastewaters is a very complex and dynamic system, where micro‐ biological and physico-chemical aspects are strongly linked. This is the reason why granular anaerobic sludge is often applied in various treatment processes, allowing higher loading rates in comparison to conventional systems with dispersed sludge. One of the examples is also the upflow anaerobic sludge blanket reactor (USBR), where wastewater is flowing through a dense bed of sludge with high microbial activity. Granules, which are formed due to the natural selfimmobilization of anaerobic bacteria, have a diameter of 1-4 mm. The system could be affected by the presence of suspended and colloidal components of the influent, such as fats, proteins, or cellulose, but these components are usually not typical for landfill leachates. USBR system is well known by its high biomass concentration, high organic loading rates and short hydraulic retention times, a lack of bed clogging, low mass transfer resistance, and large surface area. Another version of the USBR reactor is the expanded granular sludge bed reactor (EGSB) with a high upflow liquid velocity above 4 m s-1 and a large height/diameter ratio (> 20) to intensify mixing [74]. Some of the typical treatment systems with prevailing anaerobic conditions, if not completely anaerobic, are presented in Table 6. The systems are in design more or less similar to aerobic ones, presented in Table 5. Anaerobic rotating biological reactors (Table 5, Table 6) are comparable to aerobic ones; they are only covered to avoid contact with air. In both systems there is a series of rotating discs, partly or completely immersed in a reactor through waste‐ water flows. The system is not energy demanding and it is able to deal with a wide range of flows [74].

Perspectives on Biological Treatment of Sanitary Landfill Leachate http://dx.doi.org/10.5772/60924

Systems with suspended-growth biomass

Systems with attached/immobilized biomass

Activated sludge (AS):

Filters:

⋅ Continuous flow reactors

⋅ Upflow system

⋅ Sequencing batch reactors (SBR)

⋅ Downflow system

Membrane bioreactors (MBR):

Fluidized bed:

⋅ External membrane module

⋅ Sand carrier of biomass

⋅ Submerged/Immersed membrane module

⋅ Activated sludge carrier of biomass Upflow anaerobic sludge blanket (UASB): ⋅ Expanded granular sludge bed (EGSB) Moving bed bioreactors (MBBR) ⋅ Rotating biological contactors

Table 6. Typical treatment system with anaerobic microorganisms in landfill leachate treatment [2, 74, 86].

A possibility of biological treatment in an anaerobic submerged membrane bioreactor was studied in reference [87]. The treatment efficiency under different feeding conditions with different dilution rates of the stabilized leachate and synthetic wastewater (5-75 v/v%) was studied. It contained 2,800-5,000 mg L-1, 1,950-3,650 mg L-1, and 751-840 mg L-1 of COD, chloride, and ammonium, respectively. The capacity of the reactor was 29 L and it contained submerged membrane bioreactor with the capillary ultrafiltration module. Reactor was fed with granular sludge, obtained from industrial wastewater treatment plants, and experiments were carried out at 35°C. The effluent form anaerobic reactor was then further treated using reverse osmosis. Treatment was the most efficient at 20 v/v.% of landfill leachate and the system was able to remove up to 90% of COD. For leachate concentration above 30 v/v.%, significant decrease of anaerobic treatment efficiency was observed, probably due to the toxicity of the landfill leachate. The combination of anaerobic and aerobic reactors was employed in reference [88]. Here, the anaerobic sequencing batch reactor (ASBR) and the pulsed sequencing batch reactor (PSBR), both with 10 L of working volume, were combined to enhance COD and nitrogen removal from the fresh landfill leachate. Anaerobic and aerobic activated sludges from wastewater treatment plants were used to inoculate ASBR and PSBR, respectively. In ASBR, the organics from raw leachate were mainly degraded. During the 157 days long joint operation period, 89.6%-96.7% of COD and 97.0%-98.8% of total nitrogen (TN) removal were achieved. In the effluent, COD and TN were less than 910 mg L-1 and 40 mg L-1, respectively, without any extra carbon source addition. Most of the organics in the raw leachate were used as the carbon source during denitrification. In addition, excess organic polymers such as polyhydroxybutyrate (PHB) and glycogen, stored in biomass, acted as the internal carbon source during endogenous denitritation, confirming the possibility of nitrogen removal without the addition of an extra carbon source. These systems are recently more and more often applied, they stop nitrification at the nitrite stage (nitritation), followed directly by reduction to N2 in anoxic conditions with carbon addition (denitritation). Nitritation/denitritation is attractive because it reduces up to

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25% of the total oxygen requirements at the wastewater tretment plant and thus it could significantly reduce costs. Another reactor system with the up-flow anaerobic sludge bed (UASB) reactor (working volume 3 L) and a 9-L sequencing batch reactor (SBR) in series was used to treat the landfill leachate, in order to enhance the organics and nitrogen removal [89]. The UASB reactor was inoculated with the anaerobic granulated sludge from the methanogenic reactor at wastewater treatment plant, while the aerobic activated sludge from the wastewater treatment plant was used to seed the SBR. Inhibition of the free ammonia on nitrite-oxidizing bacteria and process control were used to achieve the nitrite pathway in the SBR. During a 623 day long experiment, the maximum organic removal rate in the UASB and the maximum ammonium oxidization rate in the SBR were 12.7 kgCOD m-3 d-1 and 0.96 kgN m-3 d-1, respectively. COD, TN, and NH4+N removal efficiencies were 93.5%, 99.5%, and 99.1%, respectively. In the SBR, the nitrite pathway was initiated at low temperatures (14.0°C-18.2°C) and was maintained for 142 days at temperatures 9.0-15°C. Here, stable nitritation was predominantly done by the ammoniaoxidizing bacteria. An anaerobic pilot-scale sequence batch biofilm reactor (AnSBBR) at room temperature to treat stabilized leachate from a 12-year-old landfill with two extensions, 2 and 5 years, respectively, was used in reference [90]. Leachate was collected from these two exten‐ sions. Its COD was 8,566±2,662 mg L-1, with pH around 7.95. The volume of the reactor was 746 L. It was filled with foam cubes (4 x 4 cm with density of 23 g L-1) as inert biomass support. 110 L of the biomass obtained from the existing stabilization pond was used as inoculum. The reaction time was in a range of 5-7 days with filling time of 15 minutes, while 30 minutes were used for the emptying of the system. The treatment efficiency reached over 70% of COD. The authors also studied the kinetics of the process and confirmed that the AnSBR reactor can be considered as a good alternative for the pretreatment of landfill leachate, if it is good or at least partially biodegradable. It can be concluded that the selection of the biological treatment of the landfill leachate is dependent upon many factors and that the techniques, developed at particular site, could not be always efficiently applied elsewhere [91]. This is also one of the reasons for the intensive development of novel concepts, where fungal treatment seems to be one of viable options. 4.2. Fungal treatment In the last few decades, the white rot fungi have showed great potential for the removal of hazardous and toxic pollutants. They produce various extracellular ligninolytic enzymes, including laccase (Lac) and manganese peroxidase (MnP), which are involved in the degra‐ dation of lignin and their natural lignocellulosic substrates. However, these enzymes are even capable to degrade various pollutants such as phenols, pesticides, polychlorinated biphenyls, chlorinated insecticides, organic dyes, and a range of other compounds. They have been mostly applied for treatment of textile wastewaters due to their excellent decolorization and detoxi‐ fication effect [14]. A few years ago, a successful treatment of a young landfill leachate with different strains of white rot fungi was presented [14], while the fungal treatment of leachate generated in old landfills has not been investigated so far.

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4.2.1. Treatment under fungal growth conditions In general, two strategies can be applied in water pollutant degradation: (i) direct degradation by active biomass in one reactor or (ii) use of extracted enzymes from the culture medium. In the first case, the fungal growth and the enzyme synthesis, as well as the enzymatic degrada‐ tion of the pollutant take place in the water solution of the pollutant that is in the wastewater. Here, the effect of the pollutant on the microbial growth and the enzyme synthesis must be taken into account. Fungal enzymes can be intracellular or extracellular products, synthesized during growth or after the growth phase. The pollutant depletion from the wastewater can happen due to the enzymatic degradation or only due to the adsorption on the biomass. Under aerobic conditions, aeration of the reactor is necessary; while under anaerobic conditions, methane is produced. Therefore, the reactor for the first strategy should be considered as a gas-liquid-solid phase system with proper mixing and aeration, since microbial biomass, especially when it is immobilized, can be treated as a solid phase [92]. 4.2.2. Treatment with fungal enzymes From the reactor design point of view, a much simpler configuration, which is only a liquid phase system with proper mixing, can be applied in the second strategy, when only enzyme is to be added to the wastewater. Here, the effect of the pollutant on the enzyme inhibition should be considered [92]. 4.2.3. Factors affecting fungal activities The fungal growth and enzyme production are influenced by numerous factors. Media composition has enormous effect on the fungal growth and the production of their degradation systems. In general, a special attention has to be focused on the carbon and nitrogen sources, together with mineral nutrients and other additives. The composition of landfill leachates varies with location and especially with age and is consequently a result of aerobic and later anaerobic conditions in the main body of the landfill. A carbon source is necessary for the growth and enzyme production. In the research during fungal cultivation studies, organic compounds such as glucose, sucrose, starch, and similar have been used. A young landfill leachate usually contains highly biodegradable volatile fatty acids, while in an old landfill leachate, refractory humic and fulvic acids are present, and biodegradable carbon source in various forms must be added to allow fungal growth and enzyme production. A landfill leachate usually also contains toxic phenols and xenobiotics. The white rot fungi can use inorganic as well as organic nitrogen sources. Nitrogen demands for growth and especially enzyme production differ markedly between fungal species. In the case of P.chrysosporium, production of ligninolytic enzymes is more effective under the conditions of nitrogen limitation, while B.adusta produces more LiP and MnP in nitrogensufficient media. A landfill leachate usually contains higher concentrations of inhibitory ammonium nitrogen, which must be considered during research. All microorganisms have certain requirements for other medial components, such as mineral nutrients. For example, the white rot fungi need iron, copper, and manganese. However, besides the mentioned metals,

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a landfill leachate usually contains excess concentrations of toxic heavy metals, which suppress microbial activities [93]. The majority of the filamentous fungi, along with the white rots, grows and produces enzymes optimally at acidic pH values. However, one must distinguish between the optimum pH for growth and enzyme production, the optimum pH for the action of isolated enzymes, and the optimal pH for pollutant degradation [93]. Temperature has to be considered from its influence on the growth and enzyme production, the enzymatic action and the temperature of the waste stream. Most white rot fungi are mesophiles with the optimal cultivation temperature of 27°C-30°C, while optimal temperatures for enzyme reactions are usually higher, but below 65°C [14, 93]. Ligninolytic fungi are obligate aerobes and therefore need oxygen for growth and ligninolytic enzyme synthesis. The oxygen demand depends on the fungus and their ligninolytic system. In addition, leachate treatment also requires oxygen. The problematic of the oxygen supply to the culture media during cultivation has been covered in numerous papers. The major problem is the low oxygen solubility in water (only 8 mg L-1 at 20°C). To enhance the oxygen gas-liquid mass transfer and to satisfy the microbial oxygen requirements during cultivation, aeration and agitation are necessary. This can affect the fungal morphology and lead to the decreased rate of enzyme synthesis [94]. As a result of this, various types of bioreactors have been designed, generally divided into static and agitated configurations. The choice of the reactor depends on the particular system, although the appropriate gentle agitation gives as good or even better results as those from static conditions [14, 95]. It is important to optimize an initial leachate concentration for successful degradation in terms of BOD, COD, and ammonium nitrogen removal, as well as detoxification. Namely, leachates are usually toxic to the microorganisms, while the toxicity depends on the type and age of the leachate. This problem can be easily solved by mixing with other types of the wastewaters, mainly the municipal sewage, as already discussed in Chapter 3.2. 4.2.4. Reactors for fungal treatment A variety of reactor configurations can be used, similar to those for fungal cultivation under submerged conditions. In general, reactors may be divided into two general groups, relying on the situation with the microbial cells. With the first type, the cells are kept in suspension, freely or attached to some support, where they grow and perform their metabolic activities. Suspension can be maintained by a mechanical device, such as a stirrer, through the sparging of air or both. Gentle mixing and aeration have usually been the necessary prerequisites for a successful biomass growth and enzyme production. These reactors are mainly used for the production of biomass or metabolites in various brands of biotechnology, and also applied as activated sludge systems in wastewater treatment technologies. On the other hand, there are many kinds of reactors in the group, where the cells are immobilized. Here, the cells are attached on some type of support, which is in a fixed position within the reactor volume. In these reactors, the attached and active biomass is used for a certain purpose, such as removal of the pollutants from the wastewater. It is shown that batch and continuous operations are

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effective, both having advantages and disadvantages. Reports in several papers have demon‐ strated that the repeated use of mycelia over several cycles of decolorization can last from several weeks to a few months [14, 95]. Most of the studies were done under aseptic conditions, while some were effective also during non-aseptic conditions. Toxicity of the leachate strongly affects the degradation and decolor‐ ization. For a successful design of a process with the given capacity, the reactor shape and size as well as its operation mode must be selected, and operating conditions, such as concentra‐ tions, flow rates, temperature, and pH, must be defined. For the calculation of the necessary pollutant conversion, the degradation reaction rate must be estimated on the basis of literature data or laboratory experiments. This allows for the calculation of time in the batch mode or flow throughput in continuous mode. Typical bioreactors for bioremediation purposes are presented in Table 7 [96, 97]. Bioreactors with suspended-growth biomass

Bioreactors with attached/immobilized biomass

Stirred tank

Packed bed

Bubble column

Trickling filter

Fluidized bed (activated sludge)

Fixed film Rotating biological contactor

Table 7. Typical fungal bioreactors for removal of pollutants [96, 97].

4.2.5. Review of landfill leachate treatment studies with fungi Various experiments were done in different types of cultivation vessels. Laboratory equip‐ ment, such as multiwell plates and Erlenmeyer flasks, as well as packed bed aerated columns with immobilized culture, was used in the experiments. The aim was mainly to reduce COD, color, and toxicity with various fungal cultures, mostly under sterile conditions. However, the experiments were done with pure and diluted landfill leachates, while the effects of ammoni‐ um-N and additional substrate were also investigated. The experiments and their results are presented in Table 8 and briefly discussed in the paragraphs that follow. Four ecotoxicological assays were used to asses the toxicity of a raw landfill leachate and a mixture of a raw landfill leachate and effluent coming from traditional wastewater treatment plant. All tested organism indicated that both samples exceeded the legal threshold value, showing the ineffectiveness of activated sludge treatment in the reduction of toxicity. To investigate the potential of fungal enzymes for bioremediation, the autochthonous mycoflora of the two samples was evaluated by filtration, because they are already acclimatized to their toxic environment. Ascomycetes showed to be the dominant fraction in both samples, followed by basidiomycetes. However, the samples may represent a risk for human health, since some emerging pathogens were present. A decolorization screening with autochthonous fungi was done with both samples in the presence or absence of glucose. Eleven fungi basidiomycetes and ascomycetes gave promising mycoremediation results by achieving up to 38% decolori‐ zation yields [98].

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A biological process of landfill leachate treatment with the selected fungal strains on the basis of experiments under sterile conditions in Erlenmeyer flasks was developed. Samples with different concentrations (10%-100%) of young landfill leachates (LFL) with high contents of organic matter, ammonia, salts, heavy metals, phenols and hydrocarbons were treated with Trametes trogii, Phanerochaete chrysosporium, Lentinus tigrinus and Aspergillus niger. COD removal efficiencies for P. chrysosporium, T. trogii and L. tigrinus were of 68%, 79% and 90%, respectively, with a two-fold diluted LFL. COD reductions were accompanied by high-enzyme secretion and toxicity reduction, expressed as percent bioluminescence inhibition (< 20%). The effluent was toxic to these strains at LFL concentrations higher than 50% and caused growth inhibition. On the other hand, A. niger showed to tolerate raw LFL, since it grew in this media, compared to other tested strains. However, this strain is inefficient in removing phenols and hydrocarbons since toxicity reduction was very low [15]. Landfill leachates, containing high concentrations of phenols and hydrocarbons as well as ammonium nitrogen, and exhibiting high COD and high toxicity, were treated in laboratory experiments by selected strains of white rot fungi (Trametes trogii and Phanerochaete chrysospo‐ rium), with the aim of landfill leachate detoxification. Since high amounts of ammonium nitrogen are problematic for microorganisms, its effect on mycelia growth and enzyme secretion was studied first by adding NH4Cl to the growth media. Results obtained during the 14 days of cultivation showed that ammonium chloride was not problematic below 2 g L-1, while at 5 g L-1, inhibition of growth and enzyme secretion occurred. Experiments with T. trogii and P. chrysosporium in 50 v/v.% of the leachate showed 79% and 68% COD removal efficiencies. High enzyme secretion for each strain and a reduction in phenols and hydrocar‐ bons concentrations occurred and consequently, important reduction in the toxicity was achieved [99]. Biotreatability of mature municipal landfill leachate with white rot fungus and its extracellular ligninolytic enzymes was studied in Erlenmeyer flasks. Leachates were obtained from one active and one closed regional municipal landfill, which operated for many years. Both leachates were polluted by organic and inorganic compounds. The white rot fungus Dichomitus squalens was able to grow in the mature leachate from the closed landfill. Since it utilized present organic matter as a carbon source, the results showed 60% of DOC and COD removal and decreased toxicity to the bacterium Aliivibrio fischeri. In the leachate from the active landfill, the growth of the fungus was inhibited. However, when the leachate was treated with crude enzyme filtrate containing extracellular ligninolytic enzymes, 61% and 44% removal of COD and DOC was reached. In addition, a significant detoxification was proved with the bacterium Aliivibrio fischeri and plant Sinapis alba. Results showed that the fungal and enzymatic treatment is a promising biological approach for the treatment of mature landfill leachates, so this type of research should continue [70]. Color removal of a landfill leachate in a continuous experiment by immobilized Trametes versicolor on polyurethane foam was studied in a 3 L aerated column. Initially, batch experi‐ ments in Erlenmeyer flasks under sterile conditions were conducted to find the optimum, pH, optimum co-substrate dose, leachate dilution and contact time for microbial growth, and color removal. The same immobilized fungi under optimal conditions were used after 4 and 15 days

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of initial growth in several 5-day cycles to test the reuse of fungi. Experiments with the diluted leachate were done to check the effect of organic loading on color removal. Glucose was used as a co-substrate. Results show that that the same immobilized fungi can be reused for at least four cycles, each 5 days long. By using 4-day initially grown biomass, the dilution of leachate did not significantly increase the color removal efficiency without additional glucose. How‐ ever, about 50% better results in color, BOD and COD removal were obtained by addition of 3 g L-1 glucose to the concentrated leachate. Using 15-day initially grown biomass, slightly better color removal was achieved with concentrated leachate. About a twofold increase in color removal in five times diluted leachate was achieved with this biomass, when glucose was added. Results show that T. versicolor is capable of treating a highly contaminated landfill leachate and can be considered as a potentially useful microorganism [100]. Trametes versicolor and Flavodon flavus immobilized on PUF cubes were used for treatment of landfill leachate. Experiments were done in shaken flasks at room temperature. Effects of pH and co-substrates (glucose, corn starch, and cassava) were investigated at different contact times. Treatment efficiency was evaluated based on color, BOD, and COD removal. For both types of fungi, the optimum pH was 4, the optimum co-substrate concentration was 3 g L-1 and the optimum contact time was 10 days. Depending on co-substrate and fungus, the addition of co-substrate at optimum conditions could remove 60-78% of color and reduce 52%-69% of BOD and COD. Promising results prove a potential usefulness of the white rot fungi in the treatment of landfill leachate [101]. Treatment of the landfill leachate by mycelia of Ganodermaaustrale immobilized on Ecomat (organic fibers made from oil palm empty fruit brunches), packed in a 0.63 L column, was investigated. Continuous recycling of 1 L of raw leachate and 50 v/v.% diluted leachate at a constant flow (20 mL min-1) was operated for 10 cycles. The results were evaluated on the basis of pH and COD, BOD and ammonium nitrogen removal. Only slight BOD removal was achieved for the raw leachate, while no effect was observed for the diluted sample. COD removal occurred after each cycle with the diluted leachate. Higher COD removal (51%) was observed with the diluted leachate, compared to the raw leachate (23%), after the tenth cycle. Ammonium nitrogen was also reduced after cycle 8 for the diluted (46%) and the raw leachate (31%). The results suggest that the white rot fungus G. australe can be considered as a potential candidate in the landfill leachate treatment [102]. Combined fungal and bacterial treatment was also studied. A sequential process in two 10 L laboratory reactors was carried out using a fungal sp. (Phanerochaete sp.), followed by a bacterial sp. (Pseudomonas sp.) for the degradation and detoxification of contaminants in the landfill leachate. The optimization of cultivation and process parameters for individual fungal and bacterial isolates was done with Box-Behnken design (BBD) and response surface methodology (RSM) for three variables (C source, N source and duration), considering two responses (% of COD and color removal). Treatment in the sequential bioreactor under optimized conditions removed 76.9% of COD and 45.4% of color. In addition, no statistically significant DNA damage of the cells at the end of the treatment was observed, allowing the effluent to be discharged [103].

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Type of

Volume

Organism

reactor Multiwell

2.5 mL

plates Erlenmeyer

autochtonous

Type of

Measured

Removal

leachate

Duration

parameters

efficiency

Reference

Data not given 20 days

Color

up to 38%

[98]

Young

Toxicity

up to 80%

[15]

COD

up to 90%

COD

up to 79%

Toxicity

up to 50%

mycoflora 100 mL

flasks

Trametes trogii,

15 days

Phanerochaete chrysosporium, Lentinus tigrinus, Aspergillus niger

Erlenmeyer

100 mL

flasks

Trametes trogii,

Data not given 14 days

Phanerochaete

[99]

chrysosporium, Erlenmeyer

100 mL

flasks

Dichomitus

From young

COD

61%

squalens

landfill

7 days

DOC,

44%

From closed

Toxicity

42%

[70]

landfill Packed bed

3L

Trametes versicolor Young

20 days

Color

up to 78%

[100]

Immobilized 100 mL

Trametes versicolor, Fresh

15 days

Color

up to 78%

[101]

culture on

Flavodon flavus

BOD

up to 69%

COD

up to 57%

COD

Raw leachate up [102]

column, immobilized culture on PFU foam

PFU foam Immobilized 600 mL

Ganoderma australe Data not given 10 cycles

culture

(50 min

to 23%

each)

Diluted leachate up to 52 % NH3-N

up to 46%

Table 8. Overview of some fungal landfill leachate treatment studies [15, 70, 98-102].

5. Conclusion and future perspectives Many municipal as well as industrial wastes still end at the landfill, and consequently the amount of deposited wastes is significant worldwide. Even that waste separation is increasing, organics are still in the landfill and provide good environment for microbial processes, resulting in biogas and leachate production. The composition and the amount of leachate vary with the age and are dependent upon many factors. Significant components of the leachate are organic compounds, which are degradable at the beginning of landfill operation and become more and more persistent and potentially hazardous during the biotic and abiotic processes

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in the landfill body. These toxic substances should be properly removed to prevent the environmental pollution. Besides the physico-chemical and chemical methods of leachate treatment, the biological treatment in the form of recycling and combined treatment with domestic sewage, as well as bacterial treatment with the activated sludge under aerobic and anaerobic conditions, have gained significance in the last decade. In addition, the treatment using white rot fungi and their extracellular enzymes seems to be a promising method for the removal of biodegradable and refractory organic matter from the landfill leachate. Before application to the industrial scale (scale-up), research in laboratory and pilot-plant scale are usually required from the engineering point of view. Taking this into account, laboratory toxicity and biodegradability tests are already “a must” at the beginning of the leachate treatment process for both bacterial and fungal processes. However, the activated sludge processes are more than a step forward when compared to fungal processes. From the literature review, experiments with the active sludge have been done in pilot plant reactors with volumes from several liters and a few hundreds of liters, up to lagoons with 60-80 m3 in non-sterile conditions, which are close to an economically justified situation. It can be expected that the practical application of bacterial landfill leachate treatment will increase in the near future. On the other hand, much less data from the research with fungi is available. Even if this data is promising in regards to the biodegradability tests, it is obtained mainly in the laboratory experiments, mostly in Erlen‐ meyer flasks under sterile conditions. Therefore, much more time, measured in a decade or so, will be needed for the transfer to a large scale, probably only for a particular and econom‐ ically tolerable process. A considerable amount of work has been done in the biological treatment of landfill leachate, but there is still a gap regarding mathematical modeling of this process, which has not gained in significance as it has in other fields of biotechnology. Therefore, this engineering tool should be further investigated and applied.

Author details Andreja Žgajnar Gotvajn* and Aleksander Pavko *Address all correspondence to: [email protected] University of Ljubljana, Faculty of Chemistry and Chemical Technology, Ljubljana, Slovenia

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[45] Strotmann U, Pagga U. A growth inhibition test with sewage bacteria-Results of an international ring test 1995. Chemosphere 1996;32(5):921-933. [46] Žgajnar Gotvajn A, Tratar-Pirc E, Bukovec P, Žnidaršič Plazl P. Evaluation of bio‐ treatability of ionic liquids in aerobic and anaerobic conditions. Water Science and Technology 2014;70(4):698-704. [47] ISO 9408. Water Quality-Evaluation in an aqueous medium of ultimate aerobic bio‐ degradability of organic compounds-Method by determining the oxygen demand in a closed respirometer 1999. [48] ISO 7827. Water Quality-Evaluation in an aqueous medium of ultimate aerobic bio‐ degradability of organic compounds-Method by analysis of DOC 1986. [49] ISO 9888. Water Quality- Evaluation of the elimination and biodegradability of or‐ ganic compounds in an aqueous medium-Static test (Zahn-Wellens method) 2000. [50] USEPA. Methods for Aquatic Toxicity Identification Evaluations: Phase I Toxicity Characterisation Procedures. EPA/600/6‐91/003; 1991. [51] USEPA. Methods for Aquatic Toxicity Identification Evaluations: Phase II Toxicity Identification Procedures for Samples Exhibiting Acute and Chronic Toxicity. EPA‐ 600/R‐92/080; 1993a. [52] USEPA. Methods for Aquatic Toxicity Identification Evaluations: Phase III Toxicity Confirmation Procedures for Samples Exhibiting Acute and Chronic Toxicity. EPA‐ 600/R‐92/081; 1993b. [53] USEPA. Toxicity Reduction Evaluation Guidance for Municipal Wastewater Treat‐ ment Plants. EPA 833‐B‐99‐002; 1999. [54] Isidori M, Lavorgna M, Nardelli A, Parrella A. Toxicity identification evaluation of leachates from municipal solid waste landfills: a multispecies approach, Chemo‐ sphere 2003;52:85-94. [55] Cotman M, Žgajnar Gotvajn A. Comparison of different physico-chemical methods for the removal of toxicants from landfill leachate. Journal of Hazardous Materials 2010;178(1/3):298-305. [56] Žgajnar Gotvajn A, Tišler T, Zagorc-Končan J. Comparison of different treatment strategies for industrial landfill leachate, Journal of Hazardous Material 2009;162:1446-1456. [57] Cotman M. Razvoj nove strategije za oceno vplivov ekotoksikoloških potencialov od‐ padne vode na reko Doctoral thesis 2004. University of Ljubljana, Faculty of Chemis‐ try and Chemical Technology (In Slovene). [58] Muller GT, Giacobbo A, dos Santos Chiaramonte EA, Siqueira Rodrigues MA, Mene‐ guzzi A, Bernandes AM. The effect of sanitary landfill leachate aging on the biologi‐

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[70] Kalcikova G, Babic J, Pavko A, Zgajnar Gotvajn A. Fungal and Enzymatic Treatment of Mature Municipal Landfill Leachate. Waste Management 2014;34:798-803. [71] Anastasi A, Prato B, Spina F, Tigini V, Prigione V, Varese C. Decoulorisation and de‐ toxification in the fungal treatment of textile wastewaters from dyeing processes. New Biotechnology 2011;29(1):38-45. [72] Araki H, Koga K, Inomae K, Kusuta T, Away Y. Intermittent aeration for nitrogen re‐ moval in small oxidation ditches. Water Science and Technology 1990;34:131-138. [73] Ahmed FN, Lan CQ. Treatment of landfill leachate using membrane bioreactors: A review. Desalination 2012;287:41-54. [74] Chan YJ, Chong MF, Law CL, Hassell DG. A review on anaerobic-aerobic treatment of industrial and municipal wastewater. Chemical Engineering Journal 2009;155:1-18. [75] Mehmood MK, Adetutu E, Nedwell DB, Ball AS. In situ microbial treatment of land‐ fill leachate using aerated lagoons. Bioresource Technology 2009;100:2741-2744. [76] Yalcuk A, Ugurlu A. Comparison of horizontal and vertical constructed wetland sys‐ tems for landfill leachate treatment. Bioresource Technology 2009;100:2521-2526. [77] Wei Y, Ji M, Li R, Qin F. Organic and nitrogen removal from landfill leachate in aero‐ bic granulat sludge sequening batch reactor. Waste Management 2012;32:448-455. [78] Trabelsi I, Dhifallah T, Snoussi M, Ben YA, Fumio M, Saidi N, Jedidi N. Cascade Bio‐ reactor with Submerged Biofilm for Aerobic Treatment of Tunisian Landfill Leachate. Bioresource Technology 2011;102:7700-7706. [79] Canciani R, Emrodi V, Garvaglia M, Malpei F, Passineti E, Buttigklieri G. Effect of oxygen concentration on biological nitrification and microbial kinetics ina cros-flow membrane bioreactor (MBR) and moving-bed biofilm reactor (MBBR) treting old landfill leachate. Journal of Membrane Science 2006;286 202-212. [80] Galvez A, Giusti L, Zamorano M, Ramos-Ridao AF. Stability and efficiency of bio‐ films for landfill leachate treatment. Bioresource Technology 2009;100:4895-4898. [81] Stephenson T, Pollard SJT, Cartell E. Feasibility of biological aerated filters (BAFs) for treating landfill leachate, Conference Proceedings, 9th International Waste Manage‐ ment and Landfill Symposium, S. Margarita di Pula, Cagliary, 2003. [82] Loukidou MX, Zouboulis AI. Comparison of two biological treatment processes us‐ ing attached-growth biomas for sanitary landfill leachate treatment. Environmental Pollution 2001;111:273-281. [83] Trabelsi I, Sellami I, Dhifallah T, Medhioub K, Bousselmi L, Ghrabi A. Coupling of anoxic and aerobic biological treatment of landfill leachate. Desalination 2009;246:506-513.

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[84] Kalčikova G, Vavrova M., Zagorc-Končan J, Žgajnar Gotvajn A. Evaluation of the hazardous impact of landfill leachates by toxicity and biodegradability tests. Envi‐ ronmental Technology 2011;32(2): 1345-1353. [85] Billa DM, Montalvao AF, Silva AC, Dezoti M. Ozonation of landfill leachate: Evalua‐ tion of toxicity removal and biodegradability improvement. Journal of Hazardous Materials 2005;11(2/3):235-242. [86] NTUA National Technical University of Athens. Bioreactors for metal bearing waste‐ water treatment. Available from: http://www.metal.ntua.gr/~pkousi/e-learning/ bioreactors/page_18.htm. [Accessed: 2015-01-28]. [87] Bohdziewicz J, Neczaj E, Kwarciak A. Landfill leachate treatment by means of anae‐ robic membrane bioreactor. Desalination 2008;221:559-565. [88] Zhu R, Wanga S, Li J, Wang K, Miao L, Ma B, Yongzhen Y. Biological Nitrogen Re‐ moval from Landfill Leachate Using Anaerobic-aerobic Process: Denitritation via Or‐ ganics in Raw Leachate and Intracellular Storage Polymers of Microorganisms. Bioresource Technology 2013;128:401-408. [89] Sun H, Shi, Peng Y. Advanced Treatment of Landfill Leachate Using Anaerobic-aero‐ bic Process: Organic Removal by Simultaneous Denitritation and Methanogenesis and Nitrogen Removal via Nitrite. Bioresource Technology 2015;177:337-435. [90] Contrera RC, Da Cruz Di Silva K, Morita MD, Domingues Rodrigues JA, Zaiat M, Schalch V. First-order kinetics of landfill leachate treatment in a pilot-scale anaerobic sequence by batch biofilm reactors. Journal of Environmental Management 2014;145:385-393. [91] Cortez S, Teixeira P, Oliveira R, Mota M. Landfill leachate Treatment: A Portugese case study, In: Cabral GBC, Botelho BAE, editors. Landfills Waste Management, Re‐ gional Practices and Environmental Impact. New York: Nova Science Publishers Inc; 2012; p. 83-136. [92] Pavko A. Fungal Decolourization and Degradation of Synthetic Dyes. Some Chemi‐ cal Engineering Aspects. In: Garcia Einschlag FS, editors Wastewater Treatment and Reutilization. Rijeka: InTech; 2011. p. 65-88. [93] Knapp JS, Vantoch-Wood EJ, Zhang F. Use of Wood-rotting fungi for the decoloriza‐ tion of dyes and industrial effluents. In: Gadd GM. (Ed.) Fungi in Bioremediation. Cambridge: Cambridge University Press; 2001. p. 253-261. [94] Žnidaršič P, Pavko A. (2001). The morphology of filamentous fungi in submerged cultivations as a bioprocess parameter. Food Technology and Biotechnology 2001;39:237-252. [95] Doran PM. Bioprocess Engineering Principles. San Diego: Academic Press; 1995. p. 333-391.

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[96] Martin A. Biodegradation and Bioremediation. London: Academic Press; 1999. p. 355-376. [97] Moreira MT, Feijoo G, Lema JM. Fungal Bioreactors: Applications to White-rot Fungi. Reviews in Environmental Science and Biotechnology 2003;2:247-259. [98] Tigini V, Prigione V, Varese GC. Mycological and Ecotoxicological Characterisation of Landfill Leachate before and after Traditional Treatments. Science of the Total En‐ vironment 2014;487:335-341. [99] Ellouze M, Aloui F, Sayadi S. Effect of high Ammonia Concentrations on Fungal Treatment of Tunisian Landfill Leachates. Desalination 2009;246:468-477. [100] Saetang J, Babel S. Effect of Leachate Loading Rate and Incubation Period on the Treatment Efficiency by T. versicolor Immobilized on Foam Cubes. Int. J. Environ. Sci. Tech. 2009;6(3):457-466. [101] Saetang J, Babel S. Fungi Immobilization for Landfill Leachate Treatment. Water Sci‐ ence and Technology 2010;62.6:1240-1247. [102] Noorlidah A, Wan Razarinah WAR, Noor ZM, Rosna MT. Treatment of Landfill Leachate Using Ganoderma Australe Mycelia Immobilized on Ecomat. International Journal of Environmental Science and Development 2013;4(5):483-487. [103] Ghosh P, Thakur IS. Enhanced removal of COD and color from landfill leachate in a sequential bioreactor. Bioresource Technology 2014;170:10-19.

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Wet Air Oxidation of Aqueous Wastes Antal Tungler, Erika Szabados and Arezoo M. Hosseini Additional information is available at the end of the chapter http://dx.doi.org/10.5772/60935

Abstract Wet air oxidation (WAO) is a key technology in the disposal of industrial and agricultural process wastewaters. It is often used coupled with activated sludge treatment at a wastewater treatment plant (WWTP) as preliminary conversion of toxic and/or non-biodegradable components. The process is based on a high temperature and pressure reaction of the oxidizable materials in water with air or oxygen, in most cases in a bubble column reactor. The oxidation is a chain type radical reaction. The intensification of this technology is possible with the application of homogeneous and heterogeneous catalysts, recently non-thermal radical generating methods (UV/H2O2, ozonization, Fenton type processes) gathered ground also. The most frequent use of the process is in sludge treatment and oxidation of spent caustic of refineries or ethylene plants. Keywords: Process wastewater treatment, wet air oxidation, catalysts, intensifica‐ tion of wet oxidation

1. Introduction Nowadays, many industries and some agricultural activities generate large quantities of aqueous wastes containing organic and inorganic substances known as process wastewaters (PWWs). Before discharging these wastewaters into the environment like domestic wastewa‐ ters, industrial wastes must undergo various physical, biological, chemical, or combined treatments to reduce their toxicity and to convert them into biodegradable materials.

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Liquid wastes of high organic and inorganic content have been classified as hazardous wastes and are mainly disposed of by incineration. Incineration is typically used when everything else fails and then the hazardous and polluting components are eliminated with fuel con‐ sumption, which bears high costs while producing secondary pollutants such as dioxin, furans, and sulfur dioxide. Among typical treatment methods applied for these kinds of wastes, biological treatment is the primary method for removal of organic pollutants, but often not suitable since they may contain toxic, non-biodegradable, and hazardous pollutants. Moreover, microbes are vulner‐ able to (chemical) shocks, further limiting their use in the chemical industry. Some wastewater streams are too concentrated to be cleaned effectively by biological treatment. On the other hand non-biodegradable liquid wastes generated from industrial and agricultural processes can be treated by chemical oxidation, wet oxidation, or with advanced oxidation processes (AOPs) in order to eliminate their toxicity and enhance their biodegradability. Wet oxidation, also known as wet air oxidation, refers to a process of oxidizing suspended or dissolved material in liquid phase with dissolved oxygen at elevated temperature. It is a method for treatment of waste streams that are too dilute to incinerate and too concentrated for biological treatment.

2. Thermal wet oxidation processes 2.1. Wet air oxidation Thermal wet oxidation processes use high-temperature and high-pressure air or oxygen as oxidant. Wet air oxidation (WAO) refers to a process of oxidizing wastewaters, the water containing liquid wastes under pressure with air or oxygen and at high temperature (>120°C). It is a good option for treating the high-organic content PWWs which can originate from fine chemical and pharmaceutical industries. The chemistry of wet air oxidation involves chain reactions of radicals formed from organic and inorganic compounds present in the reaction mixture. In this hydrothermal process, the organic pollutants are converted into easily biodegradable substances or completely mineralized and the inorganic compounds are converted into their form with higher oxidation value, such as sulfides into sulfates. A typical condition for wet air oxidation ranges from 180°C and pressure of 2 MPa to 315°C and 20 MPa pressure. 2.1.1. Industrial application of wet air oxidation WAO technologies have been commercialized around 60 years ago. Applying different temperatures, one can mention three different oxidation categories, namely: low temperature oxidation, medium temperature oxidation, and high temperature oxidation. Commercial application of low-temperature oxidation (100°C-200°C) involves low-pressure thermal conditioning (LPO) of activated sludge and the treatment of low-strength sulfidic

Wet Air Oxidation of Aqueous Wastes http://dx.doi.org/10.5772/60935

spent caustic. The most prevalent WAO applications are for ethylene plant spent caustic and refinery spent caustic. The ethylene plant spent caustic is traditionally oxidized at low temperatures in the range of 120°C-220°C; the main purpose is to destroy the odorous sulfide content of this effluent. The refinery spent caustic WAO application operates at 240°C-260°C that already fits the medium temperature category, as well as the WAO of some organic wastes. Other industrial wastes treated by low-temperature oxidation include cyanide and phosphorous wastes, as well as non-chlorinated pesticides. High-temperature oxidation 260°C-320°C is used for refinery spent caustic, sludge destruction, and most WAO treated industrial wastewaters. Most organic industrial wastes are oxidized in this temperature range, including pharmaceutical wastes as well as pesticides, solvents, and the complete destruction of liquid wastes of pulp and paper production and other organic sludges. 2.1.2. Zimpro® wet air oxidation and related processes The history of WAO started almost about 60 years ago, when Zimmermann observed that he could dispose pulp mill liquors using air at high pressure leading to the oxidation of organic compounds dissolved or suspended in water at relatively low temperatures in the presence of oxygen [1]. They took spent pulping liquor from a local paper mill to produce artificial vanilla flavoring (vanillin) by partial oxidation of ligno-sulfonic acids. They perfected the wet air oxidation process (or the “Zimmermann Process” as it was known), and expanded it to other applica‐ tions, including wastewater treatment. The company, developed and installed these wet oxidation plants, had a diversified history as it was established by Sterling Drug Inc. as "Zimpro" at Rothschild in 1961, building the engineering and research center along the Wisconsin River. After a long expansion period, with new developments and acquisitions Zimpro was purchased by USFilter in 1996. In April 1999, Paris-based Vivendi announced the acquisition of USFilter. Siemens bought USFilter in May 2004. Presently, it is owned by Siemens Co. under the name “Siemens Water Technologies.” Since the beginning, this process (Zimpro) had been mainly used for sewage sludge treatment, but by the early 1970s, it was applied to regenerate spent powdered active carbon from wastewater treatment processes. During 1980s, WAO began to be more useful as an industrial waste treatment technology. Zimpro Products installed the first wet oxidation unit in 1982 to treat ethylene plant spent caustic. The next year, they installed and operated a wet oxidation system in California for the treatment and detoxification of hazardous wastes. In 1992, Zimpro installed a wet oxidation system at Sterling Organics in Dudley, Northumberland, UK, for pharmaceutical wastewater treatment. Currently, about 200 full-scale WAO plants are in operation for the treatment of a variety of effluent streams (municipal sludge, night soil, carbon regeneration, acrylonitrile process effluent, metallurgical coking, ethylene production spent caustic, paper filler, industrial activated sludge, pulping liquor, warfare chemicals, paper mill sludge, explosives, monosodium glutamate production, polysulfide rubber, textile sludge, chrome tannery waste, petroleum refining spent caustic, miscellaneous industrial sludges, nuclear reprocessing wastes) [2, 3]. In 2009, 2011, and 2012, Siemens contracted with Sinopec

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to build several WAO units in PR China for the disposal of spent caustic from ethylene plants and refineries.

Figure 1. Typical flow diagram of WAO [2]

The typical wet oxidation system (Figure 1) is a continuous process using rotary compressor and pump to compress the air (or oxygen) and feed liquid stream to the required operating pressure. Heat exchangers serve to recover energy from the reactor effluent and use it to preheat the feed/air mixture entering the reactor. Auxiliary energy, usually steam, is necessary for startup and can provide trim heat if required. The residence time in the reactor vessel is several hours at a temperature that enables the oxidation reactions to proceed in some cases toward total mineralization. The reactor is a bubble column; it is coupled after the heat exchanger with a separator for the separation of the effluent and the off-gases. Since the oxidation reactions are exothermic, sufficient energy may be released in the reactor to allow the wet oxidation system to operate without any additional heat input at or above COD > ~10000 mg/L. The typical reactions during WAO: Organics + O 2 ® CO 2 + H 2 O + RCOOH* Sulfur Species + O 2 ® SO 4 2Organic Cl + O 2 ® Cl - + CO 2 +RCOOH*

Organic N + O 2 ® NH 3 + CO 2 + RCOOH* Phosphorus + O 2 ® PO 4 3-

Wet Air Oxidation of Aqueous Wastes http://dx.doi.org/10.5772/60935

The oxidation reactions occur at temperatures of 150°C to 320°C and at pressures from 10 bar to 220 bar. The required operating temperature is determined by the treatment objectives. Higher temperatures require higher pressure to maintain a liquid phase in the system. Typical industrial applications for the WAO process have a feed flow rate of 1 m3/h to 50 m3/h per unit, with a chemical oxygen demand (COD) from 10,000 mg/L to 150,000 mg/L (higher COD with dilution). During the decades of process development, huge amount of data were collected and pub‐ lished about the oxidation properties of individual compounds, different process wastewaters, and sludges. In WAO of ethylene spent caustic, the conversion of sulfide is nearly 100% at 200°C during 60 minutes. The high conversion of the polluting compounds in refinery spent caustic needs higher temperature of oxidation (260°C), residuals from pesticide and herbicide production need even higher (280°C). This is associated obviously with higher total pressure because of the increased vapor pressure of water and other volatiles. A good instance is, for illustrating the technical details and problems of WAO, the oxidation of ethylene plant spent caustic, which is one of best-elaborated technology among WAO processes. Such spent caustic liquor contains as major components the compounds listed in Table 1. The sulfur containing compounds are oxidized to sulfate, being present in the basic solution as sodium sulfate, the organic components are oxidized primarily to carboxylic acids, such as acetic, oxalic, formic, and propionic acid. At 200°C and 28 bar pressure only the partial oxidation of organic compounds occurs, as the forming carboxyl‐ ic acids are well biodegradable [4]. Compound

Concentration range

NaHS

0.5-6 %

Na2CO3

1-5 %

NaOH

1-4 %

NaSR

0-0.2 %

Soluble oil

50-150 ppm

TOC

50-1500 ppm

Benzene

20-100 ppm

Table 1. Composition of ethylene plant spent caustic

WAO reliability can be hampered by off-spec feed, which is affected by the upstream proc‐ essing and handling of the spent caustic. The feed of the WAO plant for treating ethylene plant spent caustic contains some reactive organic compounds (called “red oil,” main components are primarily aldehydes that form high molecular weight materials through the so-called aldol condensation reaction) that cause fouling not only in the ethylene compressor and the separation equipments but in the heat exchanger in front of the oxidation reactor serving for the preheating of the caustic feed and cooling of the effluent. The fouling is even worse in the

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presence of iron that usually forms insoluble scale, which has to be removed by chemical and/ or mechanical treatment. Chloride is dangerous because of corrosion, in spite of the use of special alloys in the WAO equipments. A special process is the so-called VerTech oxidation of sludges. This applies an underground installation, consisting of concentric tubes as heat exchangers going down to 1200 m depth to an oxidation vessel. Due to the depth of the vessel, the bottom pressure in the reactor is above 100 bar at 275°C without the need of high-pressure pumps at the surface. The building of the installation required drilling and casing technology developed by the gas and oil production industry. The operation of the system required a frequent descaling operation with nitric acid in order to preserve the efficiency of heat exchange in the concentric tubes. The output of this plant was 80 tons of sludge per day [5]. 2.1.3. Kinetic mechanism of WAO WAO of organic pollutants is generally described by a free-radical chain reaction mechanism in which the induction period to generate a minimum radical concentration is of great significance [6-9]. In the most detailed studied reaction in WAO, in the oxidation of phenol water solution during the induction period, practically no change was observed in phenol concentration. Once the critical concentration of free radical is reached, fast reaction takes place (propagation step) when almost all phenol is oxidized. It has been found that the induction period length depends on oxygen concentration, temperature, type of organic compound and if applied on the catalyst concentration [2, 10-13]. The pH also has influence on the induction period that actually is shorter for pH values of about 4, and is increasing with the increase in pH [14]. In wet oxidation, the reaction chains are thermally initiated [15]: RH ® R g + Hg

(1)

RH + O 2 ® R g + HO 2 g

(2)

R g + O 2 ® ROOg

(3)

ROOg + RH ® ROOH + R g

(4)

RH + (OHg , HO 2 g , ROOg ) ® R g + ( H 2 O, H 2 O 2 , ROOH )

(5)

H 2 O 2 ® 2 HOg

(6)

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In the mechanism, R is any organic molecule present in the reaction mixture. In addition to the organic radicals several inorganic radicals participate in the degradation such as H, HO, and HO2/O2-. Among these radicals HO is the most reactive, it reacts with aromatic molecules (Ph=phenol) in radical addition reaction with practically diffusion limited rate coefficient at all temperatures up to the supercritical value [16]. Ph + HOg ® PhHOg

(7)

The hydroxycyclohexadienyl radical formed in Reaction (7) from phenol is also highly sensitive to oxygen. In all real PWW and in the reaction mixtures of wet oxidation in stainless steel autoclaves, Fe ions are present with measurable concentration [17-19]. These ions and other transition metal ions present accelerate the decomposition of peroxide to HO radicals in Fenton-type processes. The steady state (propagation step) is then followed by the third step (termination step) characterized by a slow oxidation rate. The first step is the chain initiation, in which free radicals (R, OH, HO2) are produced by dissociation (1) and the bimolecular reaction of dissolved oxygen with the organic compound (2), which is found to be very slow at low temperatures. When the free radical R is formed, it can readily react with molecular oxygen to give peroxoradical (ROO) (3). The other reaction is the formation of hydrogen peroxide, which decomposes with metal catalysis to OH radicals (6). Finally the peroxo radical gives with the parent compound a free radical and hydroper‐ oxide (4). The OH radicals oxidize the parent compound into a free radical again (5). In the mechanism, the organic parent compound (RH) can react thus with molecular oxygen, the organic peroxyl (ROO ) (5), hydroxyl (OH ) (6) and hydroperoxyl (HO2 ) (3) radicals [20, 21]. Intermediates formation is of great importance in WAO and has been reviewed by Devlin and Harris for the oxidation of aqueous phenol with dissolved oxygen [22]. The conclusion was that, at elevated temperatures, oxygen is capable of three different oxidation reactions with the organic: (i) introducing an oxygen atom into an aromatic ring to form a dihydric phenol or quinone; (ii) attacking carbon to carbon double bonds to form carbonyl compounds; and (iii) oxidizing alcohols and carbonyl groups to form carboxylic acids. The ring compound intermediates (dihydric phenols and quinones) were formed under conditions near the stoichiometric ratio of phenol and oxygen, increasing in quantity when oxygen was in deficiency. The unsaturated acids, namely maleic and acrylic and saturated ones, namely formic, acetic and oxalic appear independently of phenol to oxygen ratio used. Malonic, propionic and succinic acids were identified only in case of deficit of oxygen. Malonic acid undergoes decarboxylation to produce acetic acid and carbon dioxide. Another interesting research was done about the kinetics of oxidation of phenol, and nine substituted phenols were investigated [10]. The process was studied in a 1 L stainless steel autoclave at temperatures in the range of 150°C-180°C and the initial phenol concentration was 200 mg/L. The oxidation reaction found to be the first order for oxygen and also first order with respect to phenolic substrates in both cases. The overall oxidation reaction rate was found

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to be kinetically controlled when the temperature was less than 195°C and the phenol con‐ centration was less than 200 mg/L. At higher temperatures (>240°C) and higher phenol concentration (>20000 mg/L) the overall oxidation reaction rate became mass transfer control‐ led.

Figure 2. Proposed reaction pathway for phenol oxidation by molecular oxygen.

2.2. Catalytic Wet Air Oxidation (CWAO) In the 80s significant need was revealed toward treatment of highly concentrated wastewaters of chemical and pharmaceutical production, as well as residual sludge [22, 23]. Aside from WAO, CW(A)O has been applied to many different model effluents, but relatively few works have been devoted to real and complex industrial wastes [2, 24-28]. In order to carry out wet oxidation under milder conditions (at lower temperature and pressure) an alternative way would be catalytic wet (air) oxidation (CW[A]O). Soluble transition metal salts (such as copper and iron salts) have been found to give significant enhancement of the reaction rate, but post-treatment is needed to separate and recycle them.

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Heterogeneous catalysts have the advantage that they can be used without the problem of separation and for continuous operation. Mixtures of metal oxides of Cu, Zn, Co, Mn, and Bi are reported to exhibit good activity, but leaching of these catalysts was detected [25, 29-32] The reaction mechanism of CWAO is thought to be similar to the mechanism of WAO, and the function of the catalyst is essential promoting the formation of free radicals. Kinetic study was carried out on the phenol oxidation by CWAO using aqueous copper nitrate as homogeneous catalyst. A kinetic model has been established based on the free radical mechanism where the electron transfer from copper to phenol was assumed to initiate the formation of free radicals and this led to propose that the formation of free radicals (PhO⋅ and PhOOO⋅) is primarily due to the electron transfer from metal to phenol. In this model, the reaction orders were found to be approximately 1.0, 0.5, and 0.5 with respect to phenol, oxygen, and copper concentrations, respectively. In order to verify the proposed kinetics, a series of CWAO experimental tests were done at 313-333 K, oxygen partial pressures of 0.6-1.9 MPa, and copper concentrations 0-13 mg L−1. The experimental data fitted well with the model [33]. With supported copper oxide catalyst in the temperature range of 120°C-160°C, with 0.6 MPa and 1.2 MPa oxygen pressures, it was found that the reaction was first order with respect to phenol concentration and half order with respect to oxygen partial pressure [34]. The composition of the most active heterogeneous catalysts in wet oxidation, namely that they are multicomponent, alludes on the validity of the Mars-van Krevelen mechanism well known in catalytic oxidation chemistry [35, 36]. 2.2.1. Industrial homogenous CWAO processes In these processes, homogeneous transition metal catalysts are used that need, however, to be separated and then recycled to the reactor or discarded. The first process to be mentioned was developed by Ciba-Geigy, which uses a copper salt as a catalyst. From the oxidized material the catalyst has to be separated as copper sulfide and recycled into the reactor, which is titanium lined. The unit works at 300°C and pressure above 100 bar. Three units that are installed in Germany and Switzerland have achieved high oxidation efficiencies (95%-99%) on chemical and pharmaceutical wastes at elevated temper‐ atures. The other one is the LOPROX process, a relatively low-temperature and low-pressure wet oxidation that was developed by Bayer AG for treatment of organic substances, especially aromatic compounds, which degrade too slow in normal biological plants or adversely affect the degradation of other substances. The disposal of aromatics is important for two reasons, they are forming significant portion of PWWs from the chemical industry. They are present in the clarified activated sludge also, which contains humic acids, these have aromatic part with chlorine substituents. It takes place in the presence of oxygen in acidic range in a multi-stage bubble column reactor under relatively mild operating conditions (temperature below 200°C, pressure 0.5-2.0 MPa) the catalyst is the combination of Fe2+ ions and quinone-generating substances. The residence time is 1-3 hours. Above COD value of 6-8 g/L the process is

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autothermal, no heat energy input is needed. A critical issue is the choice of structural materials of the reactor and of other hot parts of the system. Enameled or PTFE lined steel can be used up to 160°C, titanium and titanium-palladium alloys are applied up to 200°C because of the acidic pH. Several LOPROX plants are in operation at Bayer AG. They dispose PWWs from intermediate, dyestuff, pharmaceutical, paper and pulp production, and clarified sludge. Veolia developed the ATHOS® process for the treatment of clarified sludge at WWTPs. It works with Cu ion catalysis at 250°C and 50-60 bar pressure with pure oxygen. The reactor is perfectly mixed, because it has a circulation loop, not the usual bubble column. The heat exchangers are working with extremely high-temperature water as heating source. Such plant is working at the Brussels WWTP, in a complex line for the abatement of clarified sludge [37]. 2.2.2. Industrial heterogeneous CWAO processes The development of stable and active heterogeneous catalysts for WO of PWWs is a difficult task, as the substrates to be oxidized are diverse, the wastes are multicomponent, and severe conditions are needed for the completion of the reactions. At high-temperature and highoxygen partial pressure even at basic pH of the reaction mixture the leaching of the active component(s) of the catalyst into the water solution frequently occurs. As mentioned in a review, there are two catalytic WAO technologies that have been developed in the late ‘80s in Japan. Both processes use heterogeneous catalysts, precious metals deposited on titaniazirconia carriers. They are able to oxidize two refractory compounds namely acetic acid and ammonia also [27]. The NS-LC process uses a Pt-Pd/TiO2-ZrO2 honeycomb catalyst. Typical operating conditions are 220°C and 4 MPa pressure with space velocity = 2 hour-1, which with these operating conditions the oxidation of compounds such as phenol, formaldehyde, acetic acid, glucose, etc., exceeds 99% conversion. In the absence of a catalyst the removal efficiency would go down to 5%-50% [38]. The specialty of the reactor is the segmented gas-liquid flow, which means that the liquid plugs are sandwiched between two gas plugs, and the flow has a mass transfer increasing and solid deposition preventing effect. The other process, which is called Osaka Gas, is based on a mixture of precious and base metals on titania or titania-zirconia carriers in a form of honeycomb or sphere. This process has been applied in several industrial and urban wastes. A typical pilot plant at British Gas’s London Research station works at 250°C and pressure of 9 MPa, with 200 L/h feed of waste [39]. Kurita Company developed a process to abate ammonia with the oxidation agent nitrite in the presence of a supported Pt catalyst. The reaction temperature (170°C) is lower than in usual WO. One of the recent developments is the CALIPHOX process made by the National Institute of Chemistry of Slovenia and an engineering firm for treatment of industrial wastewaters with a metal oxide catalyst in the extruded form in a trickle-bed reactor. It is operating in relatively mild conditions (180°C, 4 MPa). The catalyst is based on the work of Pintar and Levec [40] who studied CuO- ZnO-Al2O3.

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2.2.3. Types of CWAO catalysts The objective of catalyst application, beside reaction rate enhancement, is to operate among milder conditions. The catalyst is usually a metal salt, a metal oxide, or the metal itself [Table 2.]. As we know, heterogeneous catalysts based on precious metals deposited on stable supports are less sensitive to leaching [25, 31, 41-44]. Different catalysts were applied and their effects were investigated on different catalytic wet oxidation processes in the past years. Pt and Ru on ceria and zirkonia-ceria supports were tested in the oxidation of acetic acid that was accompanied by the loss of activity [32]. In a following paper, the same authors described the reason for deactivation of the Pt catalyst, the accumulation of carbonate species on the surface [45]. Recently, activity of Ru-oxide on different oxide supports in acetic acid oxidation was reported. The mixed Zr, Ce oxide supported catalyst proved to be the most active [46]. Substrate

Phenol

Reaction

Oxidant,

condition

catalyst

Reactor type

Removal

Reference

120-160°C

Air, Cu

trickle bed reactor

Ir > Pd > MnO. 2.2.3.2. Supported metal oxides Metal oxides can be classified according to their physical-chemical properties. One of these properties is the stability of metal oxide. Metals with unstable high oxidation state oxides, such as Pt, Pd, Ru, Au, and Ag do not perform stable bulk oxides at moderate temperatures. Most of the commonly used metal oxide catalysts (Ti, V, Cr, Mn, Zn, and Al) have stable high oxidation state oxides. Fe, Co, Ni, and Pb belong to the group with intermediate stability of high oxidation state oxides [58]. Mixtures of metal oxides frequently exhibit greater activity than the single oxide. Cobalt, copper, or nickel oxide in combination with the following oxides of iron (III), platinum, palladium, or ruthenium, are reported as effective oxidation catalysts above 100°C [59]. In addition, combining two or more metal catalysts may improve non-selective catalytic activity. Metal oxides are usually applied in the form of powders and fine particles, and with this form of catalyst structure we can achieve maximum specific surface area, but the dispersion of the particles can create unsteady state. To keep the stable state of the catalyst, at the same time not losing the active phase, some porous supports can be used. Commonly, alumina and zeolites are used as support, but surface area of aluminum oxide is limited and the pore size of zeolites cannot be permeable for large size organic molecules. 2.2.3.3. Activated carbon catalysts for CWAO Another promising catalyst could be activated carbon (AC) that shows good properties as adsorbent for both organic materials and oxygen because of its porous structure and high surface area [60, 61]. Activated carbon is stable in highly acidic and basic media and it is also easy to prepare, which is why it is used as a catalyst for different reactions [62], and also as a support for other oxidation catalysts [63, 64]. Activated carbon can also catalyze the polymerization reactions in the presence of oxygen via oxidative coupling. Phenol oxidation over activated carbon in trickle bed reactor has been investigated [65, 66]. The activated carbon was found less active than metal oxide catalysts but more stable and more environmentally accepted, and of course cost-effective [65, 66]. Phenol conversion was compared using copper catalyst and activated carbon [67]. In the long run, copper catalyst was found to lose its activity due to leaching of copper phase. On the other hand, activated carbon also exhibited a continuous drop in phenol conversion, starting from

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nearly complete and finally reaching about 48%. However, the loss of activated carbon efficiency could be ascribed to its consumption during experiments; thus, the absolute activity of activated carbon remained stable during the long term. One of the recent studies was about the CWAO of paracetamol on activated carbon [68]. The CWAO of paracetamol was investigated both as a water treatment technique and as a regen‐ erative treatment of the carbon after adsorption in a sequential fixed bed process. They used three ACs as catalysts: a microporous basic AC and meso- and micro-porous acidic ACs. During the first CWAO experiment, they noticed that the adsorption capacity and catalytic performance of fresh basic activated carbons (S23 and C1) were higher than those of the fresh microporous acidic one (L27) despite its higher surface area. It seems that this situation changed after reuse, as finally L27 gave the best results after five CWAO cycles. Respirometric tests were also done with activated sludge and it was mentioned that in the studied conditions the use of CWAO enhanced the aerobic biodegradability of the effluent. The group also checked the different ageing by measuring the physico-chemical properties of activated carbons. 2.2.4. Application of CWAO in fine chemical industry CWAO has been applied to many different model effluents, but relatively few studies have been devoted to real and complex industrial wastes [23, 27-28, 69-73]. Even in literature there are very limited number of works that dealt with real complex wastewaters. Our focus here is the pharmaceutical industry that produces mixtures of liquid wastes containing water and organic solvents, aside from higher molecular weight organic and inorganic compounds with different concentrations and different pH. Treating these wastewaters needs special condi‐ tions. In a publication, the catalytic wet oxidation of wastewater originating from apramycin production was investigated with supported Ru oxide catalysts [72]. Ru catalyzed oxidation of wastewaters originating from meat processing and vegetable processing industries were also carried out [73]. Three rather detailed reviews were published concerning wet oxidation and catalytic wet oxidation [60, 74, 75]. They also mentioned the oxidation of miscellaneous organic compounds, but published no data specifically about the oxidation of pharmaceutical PWWs. In another research, the effect of CuO/Al2O3 was investigated—which was prepared by consecutive impregnation—on three different azo dyes (Methyl Orange, Direct Brown, and Direct Green), which were treated by CWAO. The relationships of decolorization extent, COD, and total organic carbon (TOC) removal in the dye solution were also investigated. The 99% of color and 70% of TOC removal in 2 h indicated that the CuO/Al2O3 catalyst had excellent catalytic activity in treating azo dyes [76]. In Table 3, the most characteristic results of CWAO are collected; the substrates tested are in most cases phenol and acetic acid. The latter is resistant in WO, but easily biodegradable in the biological denitrification.

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Catalyst

Application

Active Phase

Carrier

Substrate

Reference

Cu

Alumina

Phenol

Sandra et. al. 1974

Phenol

Kim et al. 1991

p-cresol

Mishra et. al. 1993

chlorophenols

Sanger et. al. 1992

Cu

Alumina Silica

Cu-Zn-Al oxide

Alumina

phenol compounds

Pintar et al. 1992

Cu-Mg-La

Zn aluminate

acetic acid

Box et.al. 1974,

Mn

Alumina

phenol

Sandra et. al. 1974

Sr115

chlorophenols

Sanger et. al. 1992

Mn-Ce

None

polyethylene-glycol

Imamura et al.1986

Mn-Zn-Cr

None

industrial wastes

Moses et al. 1954

Cu-Co-Ti-Al

Cement

phenol

Schmidt et. al. 1990

Co

None

alcohols, amines, etc.

Ito et. al. 1989

Co-Bi

None

acetic acid

Imamura et. al.1982

Co-Ce

None

ammonia

Imamura et. al. 1985

Fe

Silica

chlorophenols

Sanger et. al. 1992

Ru

Cerium oxide

alcohols, phenols, etc

Imamura et. al. 1988

Ru-Rh

Alumina

wet oxidized sludge

Takahasi et. al. 1991

Pt-Pd

Titania-zirconia

industrial wastes

Ishii et. al. 1991

Ru

Titania-zirconia

industrial wastes sludge

Harada et. al. 1993

Pt

Alumina

phenol

Hamoudi et. al. 1998

Mn

Cerium oxide

phenol

Hamoudi et. al. 1998

Ru

Titania

phenol

Vaidya et. al. 2002

Pt, Ru

Carbon black composite Silica-

phenol

Cybulski et. al. 2004

Levec et. al. 1976

titania Ru

Pelletized cerium oxide Zirconia phenol

Wang et al. 2008

Ru oxide

Titania

acetic acid

Wang et. al. 2008

Zirconia Titania-cerium oxide Zirconiacerium oxide Pt, Ru

Cerium oxide, Zr-(Ce-Pr)-O2

phenol

Keav et. al. 2010

Ru, Ru-Ce

Alumina

isopropyl alcohol

Yu et. al. 2011

phenol acetic acid DMF Graphene oxide (GO)

None

phenol

Yang et. al. 2014

Ru, Pt

Zirconia

acetic acid

Lafaye et. al. 2015

phenol

Espinosa de los Monteros et.

Cerium oxide Titania Ru, Pt

Titania Cerium oxide Titania-cerium oxide

Table 3. Summary of reported heterogeneous catalytic WO research [38, 57, 77-100]

al. 2015

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2.3. Intensification of WAO with combined technologies; effect of high energy radiation on wet oxidation at elevated temperature (combination of the two methods) As it is indicated in the previous section, one of the new areas in treating liquid wastes of highorganic content generated by the fine chemical mainly the pharmaceutical industry is the combination of AOPs with WAO. This area of work is rather new and there are very few articles dealing with this area of research. We already know that the organic compounds undergo chain type oxidation in a reactor at high temperature and high pressure, of course sometimes in the presence of a catalyst. The high temperature is needed for initiation of oxidation processes, but at the same time this high temperature and pressure could cause serious corrosion even at alkaline pH, so it could be a good idea to decrease the temperature yet induce the chain initiation of oxidation by radiation. For this reason, WO and irradiation could be combined. For radiation processing of polluted water, high-energy electron beam, γ -rays, or x-rays can be principally used. For producing the electron beam, electron accelerator devices are far the best radiation sources with respect to their output power and practical applicability. 2.3.1. Water radiolysis In the interaction between ionizing radiation (high-energy electron beam, gamma rays) and water, electronically excited and ionized molecules are formed and the product of it will be primary species such as ⋅OH, e-aq, H⋅, and molecular products such as H2, H2O2. In the presence of oxygen in water the reducing species H-atoms and the solvated electrons (e-eq) are converted into oxidizing species, perhydroxy radicals (HO2⋅) and perhydroxyde radical anions (HO2⋅). The last one together with the OH⋅ radicals are responsible for the degradation of water pollutants. H 2 O ® OHg+ eaq – + Hg+ H 2 O 2

Figure 3. Radiolysis of water.

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The radiation induced degradation of neutral phenol solution was studied in the past using end-product techniques [101-105]. There are other few works on mechanistic studies on phenol and other aromatic molecules that were carried out by combining end products and transient detections and it was suggested that the transformations were initiated by hydroxyl radical attachment to the ring and reaction of O2 with the radicals produced [106-108]. Pulse radiolysis of 2,6-dichloroaniline in dilute aqueous solution was investigated. It is known that mono- and dichloroanilines are considered to be highly hazardous pollutants in waste‐ water. These compounds are important chemical intermediates of dye and plant protection agent production. In this investigation, the hydroxyl radical formed in water radiolysis was reacted with 2,6-dichloroanliline forming hydroxy-cyclohexadienyl derivative. The irradiation was carried out at room temperature by a 60Co γ-source, built into a panorama irradiator, with 1.5 kGy h-1 dose rate. The hydroxy-cyclohexadienyl radical in the absence of dissolved O2 partly transformed to anilino radical when oxygen was present in the reaction mixture the radical transformed to peroxy radical. According to chemical oxygen demand measurements, the reaction of one OH⋅ radical induced the incorporation of 0.6 O2 into the products [109]. The irradiation-induced decolorization and degradation of aqueous solutions of azo dyes and some intermediates (anilines, phenols, triazines) were successful with electron beam irradia‐ tion. The experimental methods were the pulse radiolysis and end-product analysis with HPLC-MS. Demonstrating the practical applicability of this method, a continuous irradiation device has been built. The feed was the dye containing water of red color; the effluent was the colorless liquid during contact time of less than 0.1 second with the high energy electron beam (4 MeV) generated by a linear electron accelerator [110]. The other topic is the investigation of the degradation of pharmaceuticals that can be detected in natural waters as emergent pollutants. The radiation-induced degradation of ketoprofen in dilute aqueous solution has been tested. The intermediates and final products of ketoprofen degradation were determined in 0.4 mmol/dm3 solution by pulse radiolysis and gamma radiolysis. UV-Vis spectrophotometry and HPLC separation served for identifying the product compounds [111]. The successful degradation of organic molecules being in small concentration in waters, with high-energy irradiation generating radicals already at room temperature prompted us to combine this method with WO. Recently, this hybrid method was used and compared with the classical WO method. Noticeable conversion was observed in phenol oxidation by irradiation already at room temperature in the presence of high concentration of dissolved oxygen [112]. In one of the recent studies, WAO of highly concentrated emulsified wastewater was con‐ ducted. These kinds of wastewater usually contain all kinds of organic matters such as surfactants, additives, and mineral oils. They are typical, highly concentrated hardly biode‐ gradable organic wastewaters. This oxidation took place in a 2 L high-pressure autoclave in batch mode. The initial COD concentration of the wastewater was 48000 mg/L. After 2 h of oxidation at 220°C with supply of oxygen 1.25 times more than its theoretical value, the COD was reduced by 86.4%. The temperature seemed to be a key influential factor, especially

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between 180°C and 220°C the COD and TOC removal was evidently increased. They also recognized that with increasing the initial partial pressure of oxygen (pO2), the reaction rate significantly increased [113].

3. Conclusion The WO of PWWs became an important disposal technology, which in combination with activated sludge treatment fulfils the strict environmental requirements as well. It works at high temperature and high pressure and often has corrosive reaction mixture; therefore, its intensification is a permanent need. The most promising solutions are the catalytic wet oxidation and the combination of WO with AOP techniques, both of generate the most reactive OH∙ radicals.

Author details Antal Tungler, Erika Szabados* and Arezoo M. Hosseini *Address all correspondence to: [email protected] Hungarian Academy of Sciences Centre for Energy Research, Hungary

References [1] Zimmermann FJ, Diddams DG. The Zimmermann process and its application in the pulp and paper industry. TAPPI. 1960;43:710-715. [2] Mishra VS, Mahajani VV, Joshi JB. Wet air oxidation. Ind. Eng. Chem. Res. 1995;34:2-48. DOI: 10.1021/ie00040a001. [3] Siemens AG. “Zimpro wet air oxidation systems: The cleanest way to treat the dirti‐ est water” brochure [Internet]. Available from: http://www.energy.siemens.com/hq/ pool/hq/industries-utilities/oil-gas/portfolio/water-solution/Zimpro_Wet_Air_Oxida‐ tion.pdf [Accessed: 2015-03-05]. [4] Siemens AG, “The effects of caustic tower operations and spent caustic handling on the Zimpro wet air oxidation (WAO) of ethylene spent caustic” brochure [Internet]. Available from: http://www.energy.siemens.com/hq/pool/hq/industries-utilities/oilgas/portfolio/water-solution/Ethylene_plant_effects_on_WAO_opera‐ tions_EPC_4_2009.pdf [Accessed: 2015-03-06].

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Chapter 7

Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios Arshad Ashraf, Muhammad Saleem Pomee, Muhammad Munir Ahmad, Muhammad Yasir Waqar and Bashir Ahmad Additional information is available at the end of the chapter http://dx.doi.org/10.5772/60893

Abstract The development of a reliable decision support system and predictions for water quantity and quality often require a reasonable level of environmental and hydro‐ logical simulations at various geographic scales. The Soil and Water Assessment Tool (SWAT) model offers distributed parameter and continuous time simulation, and flexible watershed configuration and with the adoption of geographic information system (GIS) technology, a user-friendly and interactive decision support system can be developed for wastewater management. In this chapter, we evaluated the spatiotemporal evolution of wastewater contaminants in an environmentally degraded watershed through integrated field-based investigations and modeling approach. Later, management options were identified to improve the watershed health and agro-environment. The results of the modeling study exhibited variable responses of surface runoff and water quality to different scenarios of land use change. Temporal wastewater analysis indicated a significant impact of seasonality on the contaminants’ population levels. The adopted approach would prove effective in evaluating better management options to reduce negative impacts of wastewater and contaminants for sustainable agro-environment in future. Keywords: Wastewater modeling, land use scenarios, water policy

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1. Introduction Water is a scarce source in arid and semiarid areas where most of the countries face pressure due to limited opportunities to explore new water resources. This necessitates that all potential unutilized resources of water be used to increase agriculture production. The changes in surface and subsurface flows and land use conditions have direct affect on the downstream in the form of floods and/or water quality deterioration. Climate change and human interference could lead to significant spatio-temporal variations of water quantity, quality, and the associated ecological conditions besides affecting the related management systems [1]. Such complexities force researchers to develop more robust mathematical methods and tools to analyze the relevant information, simulate the related processes, assess the potential impacts/ risks, and generate sound decision alternatives. Spatially meaningful simulation of environ‐ mental flows and storages at the catchment scale is essential for predicting water quantity and quality, as well as operational management of the system [2]. There are numerous modeling wastewater efforts undertaken globally by different researchers (e.g., [1, 3–5]), the ultimate focus of which is mainly to mitigate sediment, contaminants, and non-point source nutrient; enhance water quality; and improve sustainability in agricultural production by increasing resilience. Unforeseen and undesirable consequences can result if biophysical and human systems are not examined together [6, 7]. Daloğlu et al. [4] presented a modeling framework that synthesizes social, economic, and ecological aspects of landscape change to evaluate how different agricultural policy and land tenure scenarios and land management preferences affect landscape pattern and downstream water quality. Wrede et al. [3] evaluated the performance of a fully distributed conceptual hydrologic model based on the Hydrologiska Byråns Vattenbalansavdelning (HBV) and Tracer Aided Catchment model-Distributed (TACD) model concepts in the Central Swedish lowlands. Nesmerak and Blazkova [8] employed a simple transfer function (SISO model) to describe the relationship between the daily total precipitation and the wastewater discharge at the inflow to the wastewater treatment plant (WWTP) for a large city. However, scientific quantifications were required on temporal and spatial scale to identify any feasible wastewater management solution rather than spot and one time sampling of effluents as reported by several studies (e.g., see [9–11]). The development of a sufficient understanding on which to base decisions or make predictions often requires consideration of a multitude of data of different types and with varying levels of uncertainty [12]. Wastewater contains chemicals such as nitrogen, phosphorus and levels of dissolved oxygen, as well as others that may affect its composition and pH rating. Agricul‐ tural runoff, drainage, as well as inputs from municipal and industrial wastewater often degrade the quantity and quality of surface water bodies. There is a serious need for appro‐ priate water quality monitoring for future planning and management of clean water resources. The SWAT model offers distributed parameter and continuous time simulation, and flexible watershed configuration and with the adoption of GIS technology, a user-friendly and interactive decision support system can be developed for wastewater management. The primary focus of this chapter is to assess the spatio-temporal evolution of wastewater con‐ taminants through the modeling approach and identify management options to improve the watershed health and agro-environment. The findings of the study may support policy makers, researchers, and water managers to make more robust water policy and management options under the changing environment in the future.

Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios http://dx.doi.org/10.5772/60893

2. Case study In order to develop suitable watershed management strategies, reliable investigation of the watershed problems is necessary. The influence of historical land use evolution on the yield of Rawal watershed lying in the southern Himalayan region was studied to take proactive measures to control the negative impacts of water contaminants in the downstream. The runoff from heavy rains brings a lot of sediments and wastewater from the adjoining areas that increases suspended, as well as bed load in the Korang River and ultimately in the Rawal lake. When the organic nutrients are added to the lake it causes eutrophication—as algal growth increases in the waste, dissolved oxygen concentrations are depleted and increase in sedi‐ mentation deteriorates water quality [13]. If wastewater is being discharged into the lake, then the nutrients that are of most important concern are nitrogen and phosphorus. Different studies on the Rawal watershed revealed water quality implications at the lake site. For example, biochemical oxygen demand (BOD) of about 680 mgl-1 was reported by Malik [9], while Ahmad et al. [11] reported total dissolve solids (TDS) of dam water from 131mgl-1 to 182mgl-1. The issue of dam water quality is further being aggravated by rapid unplanned urban encroachment in the Rawal watershed area since the last decade. These urban settlements are producing sewage in large quantities that ultimately drain toward the lake through freshwater streams. Since dam water supplies were mainly being used for drinking purpose, therefore, scientific investigation was required to trace the impact of urbanization on Rawal watershed runoff. Because of continuous water quantity and quality degradations primarily due to urban encroachments within the Rawal watershed, there was pressing need for its management on sustainable basis. The emphasis was to mitigate negative water quality implications particu‐ larly due to urban sewage. However, scientific quantification of waste flows over a reasonable time frame was essentially required for effective mitigation. Initially a questionnaire-based detailed survey was undertaken by selecting Bharakaho as a pilot area with specific research motivations, e.g., to investigate water consumption and wastewater disposal systems of selected urban settlement of the Rawal watershed, to quantify temporal and spatial wastewater quality and quantity of selected locations of the settlement using statistical means, and to develop recommendations for the appropriate measures of safe and cost-effective disposal of wastewater. An existing fully calibrated and verified SWAT model of the Rawal watershed [14] was used to simulate nutrient load responses under variable land use scenarios. The model calibration was performed using daily-observed flow data for a 20-year period (2001–2010). Remote sensing (RS) image data was used to analyze the agriculture and urban land use conditions and input to the model for simulating water quality parameters, e.g., sediment, nitrogen, and phosphorus loss from the landscape. 2.1. Water quality model – SWAT The biophysical water quality model—SWAT is a distributed model that integrates land management decisions with soil properties, climate information, and land topography to estimate water quality metrics at the watershed or river basin scale [15]. It is widely used for evaluating and predicting the impacts of conservation practices through simulating the effects

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of climate and land use changes on nutrient and sediment delivery from watersheds [16]. This process-based model (covering multi aspects of hydrology, soil, crop growth, nutrients, sedimentation, pesticides) divides watersheds into sub-basins and hydrologic response units (HRU) as its fundamental computational unit. Runoff flow, sediment, and nutrient loads are calculated separately for each HRU and then summed to determine the total load contribution from each sub-basin [17]. Land management decisions are represented at the HRU scale [4]. Daloğlu et al. [4] studied the impact of plausible future policy and land tenure scenarios on the delivery of available dissolved reactive phosphorus (DRP) and total phosphorus (TP) by exploring links between human and environmental systems. High surface water concentrations of nitrogen and phosphorus are correlated with inputs from fertilizers used for crops [18–20]. 2.2. Description of the study area The study area is the watershed of Rawal dam that caters water requirements of the twin cities of Islamabad and Rawalpindi, located in the Northern half of Pakistan (Figure 1). The Rawal watershed has been stretched over an area of about 272 sq km within longitudes 73º 03´ -73º 24´ E and latitudes 33° 41´ -33° 54´ N. The area falls under the scrub forest zone and supports mixture composed of Olea ferogenia (Wild Olive), Dodonea viscosa, Crissa spinarum, Acacia modesta [21]. In addition, there exist different grass species in which relative cover of Themeda anathera is maximum. It has been estimated that during an average year, the Rawal watershed area was draining about 84,000 acre-feet of runoff water through four major and 43 minor stream networks [22]. The Rawal dam was constructed over Korang river during 1960 at the toe of Rawal watershed area in Islamabad to harvest runoff water to primarily meet the drinking water requirements of the twin cities. However, land use changes within the Rawal watershed at the cost of deforestation have already affected the storage capacity of the dam. Rapid unplanned urbanization particularly in the lower valleys of the Rawal watershed over the last many years has emerged as major sustainability threat for the dam due to the contin‐ uously deteriorating water quality. According to Ghumman [23], human settlements, deforestation, pesticides, erosion, and wastes from poultry, agricultural activities, and recreational activities are the most possible reasons of contamination of the water of Rawal Lake. Untreated effluents from communal, agricultural, and poultry sectors are seriously damaging the water quality of the lake. In addition to the pollution generated by human activity, the lake also receives natural pollutants that contain the excreta of various wild animal species and fouls that enter the lake via heavy rainfall [24]. Bacteria decompose this organic matter in the presence of oxygen, thus oxygen depletion results in the Eutrophication of the lake. Similarly significant eutrophication is caused by agricultural runoff, concomitant soil erosion, and point-source discharges [25]. The land use patterns within the Rawal watershed have been changed significantly since the 1960’s and major catchment area has been deforested to accommodate the rapidly increasing urban population of Islamabad—the capital city of Pakistan—and other infrastructural develop‐ ments. During 18-year period (from 1992 to 2010), about 53% of Rawal watershed land use has been changed [26]. The changed land use features altered watershed hydrology and conse‐

Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios http://dx.doi.org/10.5772/60893

Figure 1. Location of Rawal watershed and its sub-basins.

quently the designed storage capacity of Rawal dam (47,500 acre -feet at the time of construc‐ tion) has been decreased to around 31,000 acre feet [27]. Moreover, the increased urbanization, commercial, and agricultural developments within the Rawal watershed area has also significantly deteriorated the water quality [24]. Agricultural runoff results in algal blooms, poor water clarity, and summer hypoxia (low oxygen) [28–30] that generally impact fisheries, recreation, and drinking water [31]. Bharakaho, located at about 5 km from the dam site within the Rawal watershed, was selected as a surveyed site for the present study. It is the largest urban setting in the watershed. The study sub-watershed comprises of five catchments that ultimately drain into Korang river. The selected catchments were: i) Shahdara catchment (before the bridge on Murree road), ii) Colonel Amanullah road catchment, iii) Hathala catchment, IV) Kiani road catchment and Shahdara catchment (After bridge on Murree road).

3. Material and methods In order to evaluate water quality response to varying land use changes in the Rawal watershed through hydrological modeling, satellite remote sensing data of Landsat ETM+ (Enhanced Thematic Mapper plus) of 2010 period was used in the present study.

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3.1. Preparation for model input data The image was classified into seven land use/land cover classes, e.g., conifer forest, scrub forest, agriculture land, rangeland, bare soil, built-up land, and water bodies. The image classification was undertaken using the maximum likelihood rule, which provides reliable classification results [26]. Rainfall-runoff model SWAT was calibrated for the target watershed. The statistical measures, such as coefficient of determination (R2) and Nash-Suttchiffe Simulation Efficiency (Ews), were used to evaluate model prediction. The R2 value is an indicator of strength of relationship between the observed and simulated values, while the latter coefficient indicates how well the plot of observed versus simulated value fits the 1:1 line. 3.2. Wastewater sampling strategy To complement wastewater related quantifications, a wastewater sampling study was designed by selecting Bharakaho sub-watershed due to its highest population density; rapid urbanization rate, location near to dam site, and existence of significant commercial and industrial activities, thus posing high potential of wastewater yield. Wastewater quality was monitored continuously for over 14 months in terms of eight parameters and results were analyzed in temporal and spatial context. Advanced statistical tools were employed to further investigate significance levels for temporal and spatial variability. The quantifications of wastewater pollution were performed using empirical relationships for domestic, industrial, and commercial land uses. Review of available wastewater management related options was made and based on a robustly developed criterion, the feasible option was recommended. Five catchments and their drainage pattern were delineated in Bharakaho sub-watershed using GIS tools. The sampling points were selected based upon physical surveys at outlets of these catchments (Figure 2), the particulars of which are shown in Table 1.

Location

Geographic Coordinates

Shahdara catchment (Before

73o 10ʹ 23.7128ʺ E

bridge on Murree road)

33o 44ʹ 12.0721ʺ N

Col. Amanullah Road catchment

Hathala catchment

Kiani road catchment

73o 10ʹ 23.7128ʺ E 33o 44ʹ 13.0822ʺ N 73o 11ʹ 38.9016ʺ E 33o 44ʹ 38.2142ʺ N 73o 10ʹ 42.1186ʺ E 33o 43ʹ 35.1740ʺ N

Shahdara catchment (After bridge

73o 10ʹ 37.0289ʺ E

on Murree road)

33o 43ʹ 29.6024ʺ N

Area (ha)

Total Industrial and Commercial Activities

132

3

25

7

169

20

189

19

160

11

Table 1. Location of sampling points and type of activities in the study area.

Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios http://dx.doi.org/10.5772/60893

Figure 2. Location of sampling points in Bharakaho sub-watershed.

To analyze wastewater implications, extensive review was made. The critical parameters related wastewaters associated with urbanization selected for monitoring purposes are: Biochemical Oxygen Demand (BOD); Chemical Oxygen Demand (COD); Electrical Conduc‐ tivity (EC); pH; Total Dissolve Solids (TDS); Total Phosphorous (TP); Nitrate (NO2); and Nitrite (NO3). Monthly sampling frequency was used to collect waste samples and accredited laboratory of PCRWR-Islamabad was used for analysis purpose. Statistical tools were used to test the spatial and temporal variations of wastewater in the study area. Average daily waste flows from each catchment of Bharakaho generated from different sources were estimated by using different empirical relationships as described below. 3.2.1. Domestic waste flow estimations The relationship in Eq. 1 was used for the estimation of domestic waste flow from the study area. Q DW = Pt ×q

(1)

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Where QDW = Domestic Sewage Flow (lit/day) Pt =Total Population of the Area Q = Average daily per capita water use (lit/day) 75% of water supply was assumed to be returning as sewage as suggested by Vesilind [32]. 3.2.2. Industrial waste flow estimations Industrial waste estimation was inclusive of i) waste generated per employee (17.5 GD-1) and ii) waste generated per square foot (0.18 GD-1) as recommended by the University of Minnesota [33]. Total industrial waste generation per day from industrial campus = Waste generated by employees + Waste generated from total area: WTI = WE + WTA

(2)

Where WTI = Total industrial waste generation per day from industrial campus (unit) WE = Waste generated by employees (unit) WTA = Waste generated from total area (unit) 3.2.3. Commercial waste flow estimation Commercial waste flow estimation was dependent upon type of commercial activity. For example, from hospitals, the waste flow was estimated using following formula after [33]: Total waste generated from hospitals = Waste generated by patients + Waste generated by practitioners + Waste generated from total area Waste generate per patient = 3 GD-1 The number of patients was known by visiting hospitals. Waste generated by practitioners = Waste generated per practitioner x Number of practitioners Waste generated per practitioner = 275 GD-1 [33] The number of practitioners is known by visiting hospitals. Waste generated from total area = Waste generated per square foot x Area of commercial activity Waste generated per square foot = 1.1 GD-1 [33]

3.2.3 Commercial waste flow estimation  Commercial  waste  flow  estimation  was  dependent  upon  type  of  commercial  activity.  For  example,  from  hospitals, the waste flow was estimated using following formula after [33]:  Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios Total waste generated from hospitals = Waste generated by patients + Waste generated by practitioners + Waste  http://dx.doi.org/10.5772/60893 generated from total area  Waste generate per patient = 3 GD‐1  The number of patients was known by visiting hospitals.  Area ofWaste generated by practitioners = Waste generated per practitioner x Number of practitioners  commercial activity is known through visiting the commercial activity. Similarly, Waste generated per practitioner = 275 GD‐1 [33]  wastes The number of practitioners is known by visiting hospitals.  are generated from all commercial activities. The waste is often extensive in few drains Waste generated from total area = Waste generated per square foot x Area of commercial activity  depending upon the urban development in the area (Figure 3). Waste generated per square foot = 1.1 GD‐1 [33]  Area of commercial activity is known through visiting the commercial activity. Similarly, wastes are generated  Average daily waste loading (organic and inorganic) was estimated to quantify the amount of from  all  commercial  activities.  The  waste  is  often  extensive  in  few  drains  depending  upon  the  urban  waste being generated from the study area, so that proper wastewater treatment technology development in the area (Figure 3).  

can be recommended for the area. Wastewater treatment options were analyzed and a criterion Average  daily  waste  loading  (organic  and  inorganic)  was  estimated  to  quantify  the  amount  of  waste  being  generated  from  the included study  area,  so  that  proper  wastewater  treatment and technology  can  be  recommended  for  the  was developed that costs (capital, operating, maintenance), technology, man‐ Wastewater  treatment  options  were  analyzed  and  a  criterion  was  developed  that  included  costs  (capital,  power, area.  climatic conditions, and community interactions. operating, and maintenance), technology, manpower, climatic conditions, and community interactions.  

  Figure 3. Urbanization in Bharakahu (left) and wastewater from Hathala drain entering Korang River (right). 

Figure 3. Urbanization in Bharakahu (left) and wastewater from Hathala drain entering Korang River (right).

4. Results and discussion  4.1 Water and wastewater discharge pattern  

4. Results and discussion Through  questionnaire‐based  surveys  conducted  in  the  pilot  area  of  Bharakaho,  information  related  to  water  supply,  drainage  mode,  water  consumption  patterns,  and  existing  wastewater  systems  were  gathered.  In  majority of the area (about 60%), groundwater was the major source of water supply followed by surface‐tapped  4.1. Water and wastewater discharge pattern water  from  Simly  dam  (26.3%)  for  domestic  needs  (Figure  4).  While  a  few  were  using  both  surface  and  sub‐ surface water supplies for domestic purpose (6.3%). When inquired about average water consumption, majority  Through questionnaire-based surveys conducted in the pilot area of Bharakaho, information was using 100–200 liters of water per person per day (56.60%), while significant proportion of inhabitants (21.7%)  related also reported less than 100 liter/day/person water consumption. Some also were lucky enough to have excess of  to water supply, drainage mode, water consumption patterns, and existing wastewater more than 300 liters. This trend overall suggested that water scarcity was not an issue and people were getting  systemsquite handsome amount of water supply. When inquired about waste domestic drainage mode, more than half  were gathered. In majority of the area (about 60%), groundwater was the major source of waterof the population was using buried pipe lines (62.3%), while 32.6% people were using open drains to discharge  supply followed by surface-tapped water from Simly dam (26.3%) for domestic needs wastewaters  out  of  homes.  Further,  it  was  revealed  that  almost  entire  population  was  using  septic  tanks  as  a  (Figuremeans  4). While a few were using both surface and sub-surface water supplies for domestic of  preliminary  treatment.  Further  analysis  indicated  that  combined  sewer  system  was  prevailing  in  the  (95%)  and  awareness  regarding  untreated  wastewater  discharges  into  freshwater majority streams  was was very  high  purposearea  (6.3%). When inquired about average water consumption, using 100– (77%).  While  interesting  to  note  was  the  fact  that  majority  (about  71%)  were  willing  to  pay  for  wastewater  200 liters of water per person per day (56.6%), while significant proportion of inhabitants treatment facilities if provided.

(21.7%) also reported less than 100 liter/day/person water consumption. Some also were lucky enough to have excess of more than 300 liters. 6  This trend overall suggested that water scarcity was not  an issue and people were getting quite handsome amount of water supply. When inquired about waste domestic drainage mode, more than half of the population was using buried pipe lines (62.3%), while 32.6% people were using open drains to discharge wastewaters out of homes. Further, it was revealed that almost entire population was using septic tanks as a means of preliminary treatment. Further analysis indicated that combined sewer system was prevailing in the area (95%) and awareness regarding untreated wastewater discharges into freshwater streams was very high (77%). While interesting to note was the fact that majority (about 71%) were willing to pay for wastewater treatment facilities if provided.

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Figure 4. Water consumption and discharge patterns in the studied catchments of Bharakaho.

4.2. Wastewater production potential Wastewater was not only being produced from domestic sources but there were significant industrial (construction, marble industry, and commercial activities from hospitals, markets, school, etc.) sources within the study area that were also consuming fresh water supplies and discharging wastewaters into freshwater streams. The estimated wastewater production potential per day from all these sources has been summarized in Table 2. Out of 6.354 million gallons per day (MGD) waste flow discharging from the study area, 97.4% was being added from the domestic sources, while 2.6% was contributed by industrial and commercial sources. It increases significantly at each step as river water approaches towards the lake. The Kiani road catchment was producing most non-domestic sewage flows (0.63 MGD), while it was also thickly populated area and producing highest levels of domestic sewage (2.377 MGD) followed by Hathala catchment (1.783 MGD). The wastewater sampling results are shown in Table 3. The values were averaged over 14 months of study period. The permissible limits of various wastewater parameters for Pakistan are depicted in Table 4. These limits of the parameters are National Environment Quality (NEQ) standards for wastewater parameters to be discharged in water or on land [34]. By comparing the actual parametric values with the NEQ standards, it is obvious that Shahdara (before bridge), Col. Amanullah road, and Hathala catcments were discharging BOD more than permissible levels, while COD was only exceeding for Hathala catchment and all other parameters were within

Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios http://dx.doi.org/10.5772/60893

Domestic Waste

Industrial and Commercial

Total Waste

(MGD)

Waste (MGD)

(MGD)

Shahdara catchment (Before bridge)

0.470

0.005

0.475

Col. Amanullah Road catchment

0.966

0.014

0.980

Hathala catchment

1.783

0.041

1.824

Kiani road catchment

2.377

0.063

2.440

Shahdara catchment (After bridge)

0.594

0.041

0.635

Total

6.190

0.164

6.354

Location

Table 2. Wastewater production potential from different sources in the study area.

permissible limits. However, these standards were for wastewaters, while the storage of Rawal dam was being used for drinking purposes, which required zero BOD levels, and for that reason this was unfit water. Once waste flow rates and concentrations of pollutants were estimated, the actual water loading rates in freshwater streams were calculated (Table 5). A net BOD of 2,296 kg day-1 and COD of 3,875 kg day-1 were being discharged from the area, which was considered very high. Nitrate was being discharge at highest levels (141 kg day-1), followed by phosphorus related pollutants (53 kg day-1), while TDS were being added at the rate of 19,653 kg day-1 through the study area via domestic, commercial, and industrials means (Table 5). The data indicated that spatial variability was also prevalent in the study area. For example, Col. Amanullah catch‐ ment was highly polluted due to higher values of BOD and COD (dense population and closed nature of catchment has reduced surface runoff from outside non-urbanized area), while Shahdara catchment (after bridge) was least problematic due to having large open rangeland/ vegetation cover in the catchment. According to Kahlown et al. [35] increase in population has significant effect on the water quantity and quality as the increase in population is a cause of increase in contaminants and some other wastewater parameters. BOD

COD

(ppm)

(ppm)

52

16

74

Hathala catchment Kiani road catchment

Location Shahdara catchment (before bridge) Col. Amanullah road catchment

Shahdara catchment (after bridge)

EC

TP

Nitrate

Nitrite

TDS

(μScm-1)

(ppm)

(ppm)

(ppm)

(ppm)

7.29

359

0.45

4

0.89

209

66

7.06

525

1.35

2

0.045

315

112

143

6.93

631

0.68

3

0.015

378

48

36

6.95

617

1.15

8.5

1.875

371

3

3

7.09

400

0.33

4.5

1.675

239

pH

Table 3. Average pollution loadings from the study area.

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Wastewater Parameters

Permissible Limits

Biochemical Oxygen Demand

50 mgl-1

Chemical Oxygen Demand

100 mgl-1

Electrical Conductivity

1,000 µScm-1

pH

6.0–8.0

Nitrate

20 mgl-1

Nitrite

2 mgl-1

Total Phosphates

10 mgl-1

Total Dissolved Solids

1,200 mgl-1

Table 4. Permissible limits of wastewater parameters as per NEQ standards.

BOD

COD

TP

Nitrate

Nitrite

TDS

(kg.day-1)

(kg.day-1)

(kg.day-1)

(kg.day-1)

(kg.day-1)

(kg.day-1)

Shahdara catchment (before bridge)

98

158

3

6

0.17

963

Col. Amanullah Road catchment

680

1117

11

30

0.29

2,961

Hathala catchment

547

1052

16

50

7.25

6,649

Kiani road catchment

940

1513

22

49

1.29

8,225

Locations

Shahdara catchment (after bridge)

31

35

1

6

1.73

855

Total

2,296

3,875

53

141

4.73

19,653

Table 5. Estimated loading rates of BOD, COD, TP, Nitrate, Nitrite, and TDS.

Pollution parameters such as BOD, COD, phosphates, and TDS were being discharged in large quantities (940, 1513, 22, and 8,225 kg.day-1 respectively) from Kiani road catchment, while nitrate (50 kg.day-1) and nitrite were being added (7.25 kg.day-1) from the Hathala catchment. The values of all these parameters were lowest in the Shahdara catchment. 4.3. Impact of seasonality on wastewater Careful analysis of rainfall and pollution loading rates (BOD & COD) were following reverse interaction throughout the study period. The rainfall recorded in Satrameel field station located in the study area has shown an increasing pattern during monsoon months from July to September (Figure 5). The lowest values for both BOD and COD were observed during the wet period of monsoon when increased surface runoff diluted the pollutions (least in Septem‐ ber), while higher concentrations were recorded during dry periods (highest during March) when after monsoon even base flows started decreasing (Figure 6). Although higher rainfalls during June to September 2011 had caused a lot of surface runoff that resulted in lowering of the pollution levels, nitrate and nitrite increased in amount due to the washing out of human, poultry, and animal wastes.

Modeling Wastewater Evolution and Management Options under Variable Land Use Scenarios http://dx.doi.org/10.5772/60893

to September 2011 had caused a lot of surface runoff that resulted in lowering of the pollution levels, nitrate and  nitrite increased in amount due to the washing out of human, poultry, and animal wastes.  

Figure 5. Monthly rainfall pattern during study period.

Figure 5. Monthly rainfall pattern during study period. 

  Figure 6. Temporal behavior of BOD, COD, TP, nitrite, and nitrate in the study area.  Figure 6. Temporal behavior of BOD, COD, TP, nitrite, and nitrate in the study area. To  trace  seasonality,  further  analysis  was  carried  out  and  outcome  of  BOD  and  COD  is  shown  in  Figure  7.  Seasonality  has  indicated  a  clear analysis impact  on  both  selected  during  wet is period  of  To trace seasonality, further was carried outparameters  and outcome of study  BOD during  and COD shown Monsoon  increased  freshwater  surface  runoff  in  natural  streams  caused  significant  dilution  and  consequently  in Figure 7. Seasonality has indicated a clear impact on both selected parameters during study reduced  concentrations  of  BOD  and  COD  were  observed  for  all  locations.  While  in  contrary,  during  non‐ during period Monsoon increased surface runoffcatchments.  in naturalSimilarly,  streamsthe  caused monsoon wet periods,  the of corresponding  values  were freshwater much  higher  for  respective  other  significant dilution and consequently reduced concentrations of BOD and COD were observed pollution  parameters  were  also  analyzed  in  context  of  seasonality  (Table  6).  A  close  insight  of  both  tables  revealed  that  generally  pH inincreased  with  rains  non-monsoon (monsoon)  mainly  because the runoff  waters  brought  large  for all locations. While contrary, during periods, corresponding values quantities  of  wastes,  phosphorus  and  nitrate  increased  at  the some  catchments  during  monsoon  due  to  were much higher while  for respective catchments. Similarly, other pollution parameters were increased transport of nutrients with runoff waters. EC, on the other hand, decreased during the monsoon period  also analyzed in context of seasonality (Table 6). A close insight of both tables revealed that due to dilution impact. 

generally pH increased with rains (monsoon) mainly because runoff waters brought large quantities of wastes, while phosphorus and nitrate increased at some catchments during monsoon due to increased transport of nutrients with runoff water. EC, on the other hand, decreased during the monsoon period due to dilution impact.

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Monsoon  increased  freshwater  surface  runoff  in  natural  streams  caused  significant  dilution  and  consequently  reduced  concentrations  of  BOD  and  COD  were  observed  for  all  locations.  While  in  contrary,  during  non‐ monsoon  periods,  the  corresponding  values  were  much  higher  for  respective  catchments.  Similarly,  the  other  pollution  parameters  were  also  analyzed  in  context  of  seasonality  (Table  6).  A  close  insight  of  both  tables  that  generally  pH  increased  with  rains  (monsoon)  mainly  because  runoff  waters  brought  large  192 revealed  Wastewater Treatment Engineering quantities  of  wastes,  while  phosphorus  and  nitrate  increased  at  some  catchments  during  monsoon  due  to  increased transport of nutrients with runoff waters. EC, on the other hand, decreased during the monsoon period  due to dilution impact. 

Figure 7. Spatial variability of BOD and COD during monsoon 9  and non-monsoon periods.

  EC (μScm-1) Location

Monsoon

Nonmonsoon

TP (ppm) Monsoon

Nitrate (ppm)

Nonmonsoon

Monsoon

Nonmonsoon

Nitrite (ppm) Monsoon

Nonmonsoon

Shahdara catchment

762

926

1.1

1.61

3.5

3.6

0.03

0.11

1,303

1,343

3.25

2.6

5

8.73

0.06

0.08

1,410

1,793

1.7

2.25

10.5

6.59

4.69

0.39

1,226

1,468

2.71

2.39

4.75

5.46

0.6

0.05

580

595

0.75

0.25

4.1

2.33

0.11

0.82

(before bridge) Col. Amanullah road catchment Hathala catchment Kiani road catchment Shahdara catchment (after bridge) Table 6. Impact of seasonality on EC, TP, nitrate, and nitrite.

Statistical analysis of One-Way ANOVA technique was employed to determine significance of seasonality for various parameters. BOD indicated higher values of P (>0.1) and coefficient of variance (CV=105.3) for temporal scale and lowest values (P